Wetland Drainage Effects on Prairie Water Quality Final Report Submitted by:

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Wetland Drainage Effects on Prairie Water Quality
Final Report
Agriculture Development Fund Project # 20070083
Submitted by:
Cherie J. Westbrook1, Nathalie Brunet1, Iain Phillips2 and John-Mark Davies2
1
Centre for Hydrology, University of Saskatchewan, 117 Science Place, Saskatoon, SK S7N 5C8
phone: (306) 966-1818 email: cherie.westbrook@usask.ca
2
Saskatchewan Watershed Authority, #330-350 3rd Ave. N., Saskatoon, SK S7K 2H6
Centre for Hydrology Report No. 9
January, 2011
Table of Contents
1
2
3
4
Summary ................................................................................................................................. 6
Introduction............................................................................................................................. 8
Study Site .............................................................................................................................. 12
Methods................................................................................................................................. 13
4.1
Wetland water quality characterization ........................................................................ 13
4.2
Design of wetland drainage experiment ....................................................................... 14
4.3
Natural spills and artificial ditches sampling strategy .................................................. 15
4.4
Stream water quality sampling...................................................................................... 16
4.5
Analysis of water quality .............................................................................................. 17
4.6
Ecosystem health sampling sites and frequency........................................................... 18
4.7
Macroinvertebrate assemblage sampling methods ....................................................... 19
4.8
Habitat Assessment....................................................................................................... 20
4.9
Data analyses ................................................................................................................ 20
4.9.1
Characterization of wetland water quality ............................................................ 20
4.9.2
Wetland drainage experiment ............................................................................... 21
4.9.3
Comparison of ditches and spills .......................................................................... 21
4.9.4
Stream water quality ............................................................................................. 21
4.9.5
Macroinvertebrate analysis ................................................................................... 22
5
Results................................................................................................................................... 23
5.1
Chemical characterization of wetlands ......................................................................... 23
5.2
Wetland drainage experiment ....................................................................................... 25
5.3
Comparison of ditches and spills .................................................................................. 34
5.4
Stream water quality ..................................................................................................... 35
5.5
Description of macroinvertebrate assemblages ............................................................ 40
6
Discussion ............................................................................................................................. 43
6.1
Wetland water quality characterization ........................................................................ 43
6.2
Factors controlling temporal patterns in wetland water quality.................................... 46
6.3
Water quality characteristics of a newly constructed drainage ditch............................ 48
6.4
Comparing artificial ditches and natural spills ............................................................. 50
6.5
Stream water quality ..................................................................................................... 50
6.6
Exceedance of Federal and Provincial water quality guidelines for wetlands, ditches,
spills and streams ...................................................................................................................... 51
6.7
Stream ecosystem health............................................................................................... 52
6.7.1
Effects of wetland drainage stress levels on receiving lotic ecosystem health..... 52
6.7.2
Findings on the influence of time on ecosystem health........................................ 53
6.7.3
Unanswered aspects of drainage-biotic relationships and future directions......... 53
7
Conclusions and recommendations....................................................................................... 54
7.1
Conclusions................................................................................................................... 54
7.2
Recommendations......................................................................................................... 55
8
Acknowledgements............................................................................................................... 56
9
References............................................................................................................................. 57
10
Other aspects..................................................................................................................... 65
10.1 Communications ........................................................................................................... 65
10.2 Personnel involved........................................................................................................ 65
11
Appendices........................................................................................................................ 66
2
List of Tables
Table 4.1: Summary of number of wetlands sampled in each land cover and permanence class. 14
Table 4.2: Photographs of artificial ditches (DT) and natural connections (SP) studied that drain
wetlands at Smith Creek watershed and means of physical properties measured along the
connections: discharge (Q), velocity (v), depth (d), and width (w). Sample locations were
measured from the wetland edge along the connection. ................................................... 16
Table 4.3: Wetland loss (%) and corresponding sub-basin drainage classification used in this
study. The wetland loss data are derived from wetland areal extent changes between 1958
(Ducks Unlimited Canada data) and supervised land use classification data from 2007
and 2008 SPOT5 images (Fang et al., 2010). ................................................................... 17
Table 4.4: Sites, names and coordinates included in the Smith Creek ecosystem health
assessment 2008 and 2009. ............................................................................................... 19
Table 5.1: Mean (standard error) of wetlands within each permanence and land cover class, as
well as a summary of results for two-way ANOVAs. Semiperm is semipermanent
wetlands. Differing letter superscripts indicate significantly different (α = 0.05) Tukey’s
pairwise comparisons; * and ** denote a statistically significant difference at α = 0.05
and α = 0.01, respectively. ................................................................................................ 25
Table 5.2: Mean (standard error) stream export coefficients for the 2009 freshet by site. Differing
superscripts indicate significant difference at α = 0.05, assessed using a one-way
ANOVA and the Holm-Sidak method for multiple pair-wise comparisons..................... 39
Table 5.3: Sites sampled for ecosystem health in 2008. ............................................................... 41
Table 5.4: Metric values and their probability of difference from reference (Metric Significance)
........................................................................................................................................... 42
Table 5.5: Test Site Analysis results for sites from April 13, 2009 in the Smith Creek Watershed.
D indicated the magnitude of effect or difference from reference, while ncP is the
measure of significance..................................................................................................... 43
Table 5.6: Test Site Analysis results for sites from April 13, 2009 in the Smith Creek Watershed.
D indicated the magnitude of effect or difference from reference, while ncP is the
measure of significance..................................................................................................... 43
3
List of Figures
Figure 2.1: Map of the prairie pothole region of North America (Euliss et al., 1999) ................... 8
Figure 2.2: Potential pathways of surface runoff. Runoff can a) enter streams directly; b) enter
and be stored in wetlands; stored runoff can be released from wetlands via drainage
ditches and flow either into c) other wetlands or d) streams (modified from McAllister et
al., 2000). .......................................................................................................................... 10
Figure 3.1: Historic and current day wetland and stream channel distribution in Smith Creek
watershed, produced in conjunction with Ducks Unlimited Canada................................ 13
Figure 4.1: Smith Creek watershed, Saskatchewan, Canada showing the location and ID numbers
of wetlands studied in relation to soil map units............................................................... 14
Figure 4.2: Location of the ditches (DT) and spills (SP) sampled, and the extent of open water
(includes wetlands and lakes) determined by Ducks Unlimited Canada from 2005 air
photos (courtesy of Lyle Boychuk). Inset: Air photo of LR3 wetland, showing the
sampling sites in the wetland and along the newly constructed drainage ditch (DR). The
location of the water level recorder (PT2X) is indicated by a cross................................. 15
Figure 4.3: Location of stream sampling sites and the Environment Canada hydrometric gauge
along Smith Creek. Sub-basins were identified by Fang et al. (2010). ............................ 17
Figure 4.4: Map of Smith Creek watershed with ecosystem health assessment sampling locations
from: a) 2008, and b) 2009. Note that Smith Creek at Werle Farm is TV1 and Smith
Creek East of Langenburg is SC3..................................................................................... 19
Figure 5.1: Distribution of cation and anion dominance groups as a function of specific
conductivity for the 67 wetlands studied grouped by A) land cover and B) permanence
classes. Semiperm is semipermanent wetlands................................................................. 24
Figure 5.2: Daily rainfall and volume in wetland LR3 prior to and during drainage. .................. 26
Figure 5.3: Concentration of nitrogen, phosphorus, and bacteria measured in the LR3 wetland
prior to the drainage experiment and in the newly constructed ditch. .............................. 27
Figure 5.4: Total mass of nitrogen, phosphorus, and bacteria measured in the LR3 wetland prior
to the drainage experiment and the cumulative mass exported via the newly constructed
ditch................................................................................................................................... 28
Figure 5.5: Mass of various water quality parameters, normalized to May 1, 2008, measured in
wetland LR3...................................................................................................................... 29
Figure 5.6: Mass ratio of DIN:DIP (NO3+NH4:PO4) measured in the LR3 wetland prior to its
drainage. Values below seven indicate N limiting conditions and values above 12 indicate
P limiting conditions (Wetzel, 2001). ............................................................................... 30
Figure 5.7: pH and concentration of DOC and major ions measured in the LR3 wetland prior to
the drainage experiment and in the newly constructed ditch............................................ 31
Figure 5.8: Total mass of DOC and major ions measured in wetland LR3 prior to the drainage
experiment and the cumulative mass exported via the newly constructed ditch. ............. 32
Figure 5.9: Slopes of normalized concentrations measured along the newly constructed ditch
draining the LR3 wetland at: a) 1 hr, b) 4 hr, c) 6 hr, and d) 23 hr after the start of the
drainage experiment. Concentrations were normalized by dividing the concentration at
each sample point along the ditch by the concentration at the first sampling point (DR1)
in the ditch. A value of one was then subtracted to set the intercept to zero. Only
constituent slopes that were at least marginally different (α = 0.10) from the chloride
slope and that had significant (α = 0.05) linear relationships between normalized
concentration and distance are shown. P values indicate the level of significance for
4
slopes differing from chloride, suggesting that a portion of the variability is due to biotic
processes or sorption......................................................................................................... 33
Figure 5.10: Mean and standard error of ditch and spill physical properties where * and **
denote statistically significant differences at α = 0.05 and 0.01, respectively.................. 34
Figure 5.11: Mean and standard error of A) nutrients and DOC, and B) major ions, SC, and pH in
ditches and spills where * and ** denote statistically significant differences at α = 0.05
and 0.01, respectively. ...................................................................................................... 34
Figure 5.12: Mean and standard error of ditch and spill nutrient and DOC loads where *denotes a
statistically significant difference at α = 0.05. .................................................................. 35
Figure 5.13: Discharge and water quality concentrations during the 2009 freshet at sites
characterized by differing watershed drainage: high (TV1), medium high (SC4), and low
(SC5). No data are presented prior to April 8, 2009 due to ice at the gauge station. ....... 36
Figure 5.14: Ratios of DIN (NO3 + NH4) to DIP (PO4) measured near the Smith Creek watershed
outlet (SC3) during 2008 and 2009. Values below seven indicate N limiting conditions
while those above 12 indicate P limiting conditions (Wetzel, 2001). .............................. 37
Figure 5.15: Piper plots for: basin outlet (SC3) and sub-basins characterized by high (TV1),
medium high (SC4), and low (SC5) wetland drainage. Blue and red circles indicate
samples obtained during the snowmelt period (March 23 – April 19, 2009) and following
snowmelt (April 25 – May 26, 2009), respectively. ......................................................... 38
Figure 5.16: Total exports of: a) water (Q), nitrogen (TDN, NO3, and NH4) and phosphorus (TP
and orthoP); b) major ions; and c) bacteria (E. coli and T. coli) during the 2009 freshet (3
April - 26 May) at sub-basins characterized by low, medium high, and high wetland
drainage. Exports for the SC3 subbasin (c.f. Figure 4.3) were calculated by subtraction of
exports for the three subbasins from that for the whole watershed. Total height of the bars
indicates export for the whole Smith Creek watershed. ................................................... 40
5
1 Summary
This report describes factors influencing the spatial variation in wetland water quality and
how drainage of wetlands affects downstream receiving waters in terms of their water quality
and biotic health. The specific objectives of this work were to: 1) characterize the spatial and
temporal variation in water quality of prairie potholes after snowmelt; 2) quantify solute export
along a newly constructed wetland drainage ditch; 3) characterize solute export from drained
pothole wetlands; 4) determine the extent to which stream water quality is influenced by wetland
drainage; 5) contribute to the understanding of how wetland drainage affects ecosystem health.
The research was conducted at the Smith Creek watershed, southeastern Saskatchewan, where
there has been controversy over recent renewed efforts to drain wetlands to increase agricultural
production.
A total of 67 wetlands were sampled following snowmelt in 2009 to determine whether
spatial variations in wetland water quality could be attributed to different land cover and
permanence classes. It was found that crop wetlands had greater TP and K than wood and grass
wetlands; TP, TDN, and DOC were higher in seasonal than permanent wetlands; and salts were
lower in wood compared to crop and grass wetlands. Measures of water quality of one permanent
wetland over a 20 week period in 2008 showed that the wetland acted as a trap for nutrients, salts
and bacteria. Variations in salts and DOC loads were linked to hydrological processes whereas
variations in nitrogen, phosphorus, and bacteria loads were associated with both biotic and
hydrological processes.
The permanent wetland studied was experimentally drained in November 2008. The
experiment demonstrated that wetland water quality was an important control of water quality in
drainage water. The wetland ditch acted as a simple conduit, meaning that there was little loss or
gain of solutes along the length of the ditch. Results also show that the efficiency with which a
wetland is drained and the water quality characteristics of the wetland are the factors critical to
determining solute exports via ditches. In spring 2009, water quality along seven ditches and five
natural connections that form between wetlands (termed spills) was found to be similar, except
for concentrations of TDN, DOC, HCO3, K, and Ca that were higher in ditches than spills. As in
the wetland drainage experiment, little change in the water quality along ditches and spills was
found, likely due to the low temperatures occurring in spring that can restrict biotic processing.
The important difference however was the physical characteristics of ditches and spills. The long
ditches connect wetlands to streams whereas short spills connect wetlands to other wetlands.
Thus ditches have a greater potential to contribute to downstream loading of nutrients, salts and
bacteria.
Water quality of streams draining three subbasins of the Smith Creek watershed with
varying wetland drainage intensities was compared during the 2009 freshet. Nearly all export
coefficients for the water quality parameters studied were higher in streams draining subbasins
with greater wetland drainage. Total solute export was greatest in the subbasin with medium high
wetland drainage, due in part to its comparatively large size. Although water samples generally
did not exceed guidelines for nitrogen and salts, the provincial objective for TP was frequently
exceeded. Consequently, wetland drainage is likely to exacerbate downstream eutrophication.
The macroinvertebrate-based ecosystem health of Smith Creek watershed in stream
reaches downstream of drainage activity was assessed and compared to ecosystem health at
similar sites across southern Saskatchewan. Sampling was conducted through 2008 and 2009 at
four and eight sites in each of these respective years. It was found that wetland drainage does not
6
significantly impact the ecosystem health of receiving waters, and may in fact benefit the
macroinvertebrate assemblage, at least in the short term, by increasing the amount of water and
habitat available. However, this project did not address the potential decline in ecosystem health
in source-wetlands that are being drained, where the amount of habitat and water is decreasing.
Overall, this study represents the first field-based research in the Prairie Pothole Region
on the effects of wetland drainage on downstream water quality and stream ecological health.
Therefore, results should help support effective management decisions regarding future pothole
drainage or restoration, balancing public and private costs and benefits. Several specific
recommendations were made regarding the challenges faced by this study. It is recommended
that similar studies be conducted in other geographic areas of the prairies where climate, soils,
wetland configuration and drainage may produce differing results.
7
2 Introduction
Prairie pothole wetlands are small, shallow depressions that typically lack surface water
connections and are ubiquitous across the prairie pothole region (PPR) of North America (Figure
2.1). The PPR is estimated to cover approximately 715 000 km2 (Euliss et al., 1999). Pothole
wetlands formed during the last glacial retreat that created the hummocky, undulating terrain
typical of the prairies (Tiner, 2003). About 40%, or 156 000 km2, of the PPR located in Canada
consists of hummocky moraines that have a wetland density of 18 wetlands/km2. The remaining
60% of the Canadian PPR landscape is comprised of mostly lacustrine and fluvial materials that
have, on average, 5 wetlands/km2 (National Wetlands Working Group, 1988). Pothole wetlands
provide important hydrological and ecological functions. For example, the PPR represents only
10% of the continent’s waterfowl breeding area but it produces half of North America’s
waterfowl in an average year (Smith et al., 1964; Batt et al., 1989).
Figure 2.1: Map of the prairie pothole region of North America (Euliss et al., 1999)
Many of the wetlands located in the hummocky moraine region are isolated prairie
potholes. These potholes range in permanence from those that contain water for only a few days
following spring snowmelt, to those that are continuously inundated. Most potholes do not
normally contribute to streamflow (Stichling and Blackwell, 1957). During very wet conditions
temporary surface connections can occur among them (Leibowitz and Vining, 2003; Winter and
LaBaugh, 2003), which is referred to as the ‘fill and spill’ mechanism (Spence and Woo, 2003;
Spence, 2006).
Water quality is defined differently by engineers, ecologists, hydrologists, etc., and can
encompass a wide range of water quality descriptors depending on the end user, the context of
interest, and natural conditions (Meybeck, 2005). Water quality, as defined in this report, is
based on chemical, physical, and biological descriptors that affect the structure and function of
ecosystems as well as those that negatively impact human and livestock health if in elevated
concentrations. Isolated prairie potholes exhibit high spatial and seasonal variations in water
quality (e.g., LaBaugh and Swanson, 2004). Previous studies have documented variations in
select water quality variables in relation to individual factors such as wetland permanence, water
sources, and land use of the catchment area (Rözkowska and Rözkowski, 1969; Miller et al.,
1985; LaBaugh et al., 1987; Swanson et al., 1988; Neely and Baker, 1989; Detenbeck et al.,
2002; Waiser, 2006). For example, Driver and Peden (1977) studied prairie pond chemistry in
Saskatchewan and Manitoba and found differences among temporary, semi-permanent, and
permanent ponds. Crosbie and Chow-Fraser (1999) found higher nutrient concentrations in
Ontario marshes that a greater proportion of agricultural land in their watersheds. However, a
8
concurrent assessment of nutrients and salts in potholes is needed given that different
mechanisms interact to drive their concentrations (LaBaugh et al., 1987; Wetzel, 2001). As well,
factors such as permanence and land cover type do not act in isolation on the landscape and
could instead interact with one another to regulate wetland water quality. Studies examining the
potential governing influence of the interaction of permanence and land cover type on pothole
water quality are lacking.
Approximately 40-70% of the potholes in the western prairies have been drained since
1900 to increase agricultural production (Tiner, 1984; Dahl, 1990; Brinson and Malvarez, 2002).
Recently, there have been renewed efforts to drain potholes (Watmough and Schmoll, 2007).
Temporary wetlands are the most common permanence class in the PPR, and the most impacted
by drainage and farming practices (Euliss et al., 2001). Issues surrounding pothole drainage
centre around attempts to balance the social benefits and private costs associated with potholes
on agricultural lands (Porter and van Kooten, 1993). The social benefits of potholes are believed
to be water storage and flood attenuation, wildlife habitat, and improvements of water quality
(Saskatchewan Watershed Authority, 2007a). Costs accrued by private landowners include the
nuisance of farming around potholes, delayed seeding dates, and the foregone opportunity to
increase agricultural production (Scarth, 1998; Brinson and Malvarez, 2002; Saskatchewan
Watershed Authority, 2007a). Drainage impacts remain a critical information gap for the
agricultural community, with the potential to have broad-scale economic impacts.
Drainage ditches create new surface water connections between wetlands that were
previously isolated and other wetlands, roadside ditches, and streams. The new connections
transform the hydrologic conditions of the prairies such that previously non-contributing areas
now contribute to streamflow. Figure 2.2 illustrates the different ways that surface water can
enter a stream where drainage ditches are present (McAllister et al., 2000). Isolated wetlands
have the potential to intercept and store surface runoff (Figure 2.2b). Drainage ditches can
transport water from one wetland to another (Figure 2.2c), which can cause local flood damage
to cropland or communities surrounding the terminal wetland. Drainage ditches can also
transport surface water runoff directly from a wetland to a stream (Figure 2.2d). Drainage of
many potholes has the potential to significantly increase downstream flood frequencies and
magnitudes (Campbell and Johnson, 1975). For example, Yang et al. (2008) used the SWAT
model to show that a loss of 70% of 1968 wetlands in the Broughton’s Creek watershed in
western Manitoba to drainage and degradation increased the basin’s contributing area by 31%
(19 km2), increased peak flows by 18%, and increased stream flow by 30%. Pomeroy et al.
(2009) used the newly developed Prairie Hydrological Model to show a 117% increase in
streamflow volumes with complete drainage of the wetlands in Smith Creek watershed,
Saskatchewan based on the 2007-2008 meteorological conditions. Further, a Saskatchewan
Watershed Authority (2008) assessment identified agricultural drainage as an important
contributor to high water levels in terminal (Waldsea) and near-terminal (Deadmoose, Houghton,
Fishing) lake basins in 2007.
Pothole wetlands have been shown to trap nutrients, salts and bacteria from catchment
runoff (Neely and Baker, 1989; Crumpton and Goldsborough, 1998). Wetland water quality is
thus expected to be an important control of water quality in drainage water. As wetland drainage
connects potholes to downstream water bodies, it has the potential to generate excessive nutrients
and sediment pollution that may impact the water quality, and ultimately the ecosystem health of
receiving streams and rivers (Leibowitz and Vining, 2003; Whigham and Jordan, 2003). To date,
however, there have not been field studies to support or refute this conjecture. Despite a lack of
9
data, the popularly held belief that wetland drainage will impair downstream waters has
influenced decision making in the courts. For example, a recent decision by the Water Appeal
Board (16 August 2007) in the case of Ducks Unlimited Canada vs. Jack Kalmakoff to close a
drainage ditch was due, in part, to a perceived risk of impact to downstream aquatic systems
resulting from wetland drainage. While no studies of water quality in wetland drainage ditches
has been conducted in the PPR, there is a body of literature describing nutrient export along
upland drainage ditches located in low order agricultural watersheds, which may prove useful to
understanding solute export from potholes drainage. Main findings are reviewed in the following
two paragraphs.
Figure 2.2: Potential pathways of surface runoff. Runoff can a) enter streams directly; b) enter and be stored in
wetlands; stored runoff can be released from wetlands via drainage ditches and flow either into c) other wetlands or
d) streams (modified from McAllister et al., 2000).
The water regime in upland drains is characterized by the advective flow of water
downstream. Turbulent mixing encourages oxygenation of the water. For redox-sensitive water
quality parameters, increased oxygenation during transport along ditches would be expected to
change their concentrations. Kemp and Dodds (2001) showed stimulation of nitrification and the
inhibition of denitrification in a 2nd order prairie stream with higher dissolved oxygen (DO)
concentrations, which would likely result in a net reduction in N removal along the stream
(Birgand et al., 2007). Stimulated nitrification rates can also lead to increased transport of
nitrogen (N) to receiving streams since nitrate (NO3-) is quite mobile compared to ammonium
(NH4+), which is easily adsorbed to negatively charged ditch surface particles (Strock et al.,
2007). Solute sedimentation rates vary inversely with the velocity of flowing water (Julien,
2002), affecting concentrations of water quality parameters that have an affinity for the
particulate phase, for example, phosphorus (P).
Recent studies have shown relatively high retention of mostly P (as well as some N) in
upland ditches where soils have substantial sorption or retention capacities (Sharpley et al., 2007;
Strock et al., 2007). Nguyen and Sukias (2002) showed ditch sediments in New Zealand
contained 42-57% of P originating from agricultural catchments loosely bound with Al, Fe, and
CO3-, and 6-39% of P stored more permanently in the sediment as refractory P. They also
showed the proportion of P transported was governed by the form of P and the retention
characteristics of the ditch sediments. Periodic high flow events that occur during snowmelt and
significant rainfall events increase velocity, shear force, and scour along the ditch bottom
causing the re-suspension of sediments and organic matter and consequently their downstream
10
transport (Sharpley et al., 2007; Birgand et al., 2007). Also, macrophytes and algae living in the
drains have been shown to temporarily store nutrients (Needleman et al., 2007). When they die
during season changes, water level drops, or other causes, they can contribute to the organic
matter content of the water column and also to the accretion of drain sediment. Macrophytes can
also play an indirect role in contaminant retention by reducing flow velocities and re-suspension
rates, and increasing sedimentation rates (Birgand et al., 2007).
There is very little published research on the relationship between land use and wetland
drainage on stream water quality in the Canadian Prairies. Carpenter et al. (1998) is the seminal
paper outlining the issue of how stream water quality is controlled by land use. Alberta
Agriculture and Rural Development prepares annual reports cards for stream water quality, as
measured through a water quality index, in Alberta’s watersheds with varying agricultural
intensities. Saskatchewan Watershed Authority makes similar reports for watersheds in
Saskatchewan using a different water quality index. Stream biotic integrity is commonly
assessed through measures of the aquatic macroinvertbrate community. Summaries of the
ecosystem structure and function in macroinvertebrate communities are used as proxies of
ecosystem health, and are frequently applied to describe impacts of human activities. In
particular, specific components of the macroinvertebrate assemblage are sensitive to abiotic
changes in the aquatic environment, and the expression of these changes as metrics relative to
reference, or unaffected streams, is to evaluate impact. By identifying metrics that respond to
specific stressor, management can be applied to the stressors in a watershed that are causing
ecosystem degradation.
Leibowitz and Vining (2003) suggest that stream water quality is likely to be altered by
wetland drainage as this would hydrologically connect previously isolated wetlands to streams.
Similarly, how Saskatchewan macroinvertebrate assemblages respond to stressors brought about
through wetland drainage (e.g., nutrients, sediment etc.) is poorly understood. Tangen et al.
(2003), in their attempt to create a macroinvertebrate index of biotic integrity found only a weak
correspondence of macroinvertebrate assemblages and land use in the Prairie Pothole region of
North Dakota due to an inadequate number of low impact sites. Wetland drainage was not one of
the land use factors considered. Further, as benthic macroinvertebrate assemblages are affected
by a number of factors in addition to wetland drainage, changes in assemblage composition
should currently be interpreted as a complimentary tool to primary water quality tests.
The overall objective of this research project was to determine changes in water quality
of streams and impacts to stream ecosystem function associated with wetland drainage. The
specific research objectives of this study were to:
1) characterize spatial and temporal variations in water quality of prairie potholes after
snowmelt;
2) quantify solute export along a newly constructed wetland drainage ditch;
3) compare solute export from naturally and artificially drained pothole wetlands;
4) determine how stream water quality is influenced by wetland drainage; and
5) contribute to the understanding of how wetland drainage affects ecosystem health by
quantifying benthic macroinvertebrate communities.
Saskatchewan Watershed Authority was sub-contracted to complete objective 5.
11
3 Study Site
The research was conducted at Smith Creek watershed (SCW; 50°50'4"N 101°34'48) in
southeastern Saskatchewan. The watershed is ~400 km2 (Pomeroy et al., 2009) with an effective
contributing area of ~58 km2 (Environment Canada, 2006). The contributing area is continuously
increasing as producers drain more potholes to increase arable area. Smith Creek drains into the
Assiniboine River and then Lake Winnipeg where excessive N and P loadings are causing
eutrophication (Armstrong, 2002). Snowmelt is the dominant, and often only, hydrological event.
The terrain of SCW is level to undulating and rolling. SCW is located in the Aspen
Parkland Continental Prairie Wetland sub-region (National Wetlands Working Group, 1988).
Soils in the watershed are a mixture of Black (Oxbow) and Thick Black (Yorkton) chernozems
formed in loamy glacial till (Agriculture and Agri-Food Canada, 2009).
The regional climate is semi-arid. Average seasonal temperatures are -17.9oC and 17.8oC
for the winter and summer, respectively (as measured at the Yorkton airport, ~50 km west). The
mean (1942-2009) annual precipitation is 438 mm with 28% falling as snow. Precipitation
amounts prior to the 2008 study period were in the 53rd, 34th, and 26th percentiles for winter
2007, May – October 2007, and winter 2008, respectively. Summer 2008 and winter 2009
precipitation amounts were in the 60th and 66th percentiles, respectively. Total precipitation for
the months of May 2008 (18 mm) and July 2008 (208 mm) were respectively in the 13th and 95th
percentile of values measured at the Yorkton weather station.
Land uses were determined by Pomeroy et al. (2009) using unsupervised classification of SPOT
5 images from October 1, 2008. The dominant land use is agriculture, which occupies 54% of the
watershed. Common crops include wheat, canola, and flax. Eight percent of the watershed is
grassland and pasture. Wooded areas and wetlands/open water account for 23% and 11% of the
watershed, respectively. Wooded stands are characterized by trembling aspen (Populus
tremuloides) with pockets of balsam poplar (Populus balsamifera), together with an understory
of mixed herbs and tall shrubs. Wetland vegetation is predominately willow (Salix spp.), cattails
(Typha latifolia L.), sedges (Carex spp. and Scirpus spp.), duckweed (Lemna spp.), pondweed
(Potamogeton spp.), and water smartweed (Polygonum amphibium L.). Grasslands are comprised
largely of western porcupine grass (Stipa curtiseta), plains rough fescue (Festuca hallii), pasture
sage (Artemisia frigida), and Lewis wild flax (Linum lewisii). The majority of the wetlands in
SCW belong to the marsh and shallow open water classes (National Wetlands Working Group,
1988). The average wetland density in the basin is ~20 wetlands/km2. Many of the wetlands in
the basin are typical, isolated prairie potholes that formed in glacial depressions, which at
average surface water level have no surface inflows or outflows. The wetlands range from
temporary through permanent. Recently, a number of drainage ditches have been constructed to
enhance agricultural production. These drainage ditches either completely eliminate the wetland
or drastically lower the water levels and create downstream connections with other wetlands,
roadside ditches, or streams. Between 1958 and 2001 the portion of land cover occupied by
wetlands in the Smith Creek watershed decreased ~60%%. Drainage has been most prolific in
the eastern and southwestern areas of the watershed (Figure 3.1). Many water control structures
such as road culvert gates exist in the basin and are operated by local farmers to regulate the
runoff in their cropland areas; the gates are closed during extremely high runoff periods, i.e.
during fast snowmelts or intense rain storms but remain open otherwise.
12
2009
1958
Figure 3.1: Historic and current day wetland and stream channel distribution in Smith Creek watershed, produced in
conjunction with Ducks Unlimited Canada.
4 Methods
4.1 Wetland water quality characterization
Water quality in 67 wetlands (Figure 4.1) was assessed following snowmelt in 2009.
Wetland selections were made based on obtaining relatively equal numbers of wetlands in the
different land cover and permanence classes (Table 4.1), as well as on accessibility. Land cover
classes were crop, wood, and grassland while permanence classes were seasonal, semipermanent, and permanent. Wetland permanence classes were determined using a combination
of the vegetation structure of the pothole, as per Millar (1976) and Stewart and Kantrud (1971),
assessed in May and mid-August 2009, in combination with observations of the presence or
absence of surface water recorded on air photos (26 October 1959 and ~31 May 2001), and
SPOT 5 imagery (5 July 2007). Water samples (1 L) were collected once during 19-21 May 2009
from roughly the deepest point in the wetland, which is typically the centre of the ponds, at the
midpoint in the water column.
13
Figure 4.1: Smith Creek watershed, Saskatchewan, Canada showing the location and ID numbers of wetlands
studied in relation to soil map units.
Table 4.1: Summary of number of wetlands sampled in each land cover and permanence class.
Land Cover Type
Crop
Grass
Wood
Total
Seasonal
7
9
6
22
Permanence Class
Semipermanent
6
7
7
20
Permanent
8
8
9
25
Total
21
24
22
67
4.2 Design of wetland drainage experiment
A drainage experiment was carried out on a permanent wetland, LR3. It was selected for
study because hydrological data for it prior to drainage were available from a concurrent study
(c.f. Minke et al., 2010). Additionally, the landowner was keen to drain it and the drain was
expected to connect to Smith Creek.
Water samples were collected weekly at the center of the wetland LR3 from 18 April to
22 October 2008. Water level was measured hourly using a PT2X pressure transducer
(Northwest Instrumentation Inc.) which was located near the edge of the wetland. Rainfall was
measured nearby using a tipping bucket (Texas Electronics Inc., TR-525M) and a standard
volumetric rain gauge. A crawler excavator and professional operator were employed to
construct the artificial drainage ditch (DT6) starting 19 November 2008. Water sample collection
14
began 20 November 2008 (1 hr) once the ditch was completed and connected to a downstream
wetland. Three additional sets of samples were collected 4 hrs, 6 hrs, and 23 hrs after the start of
drainage. Water samples were collected from within the wetland and at points 45 m, 70 m, 110
m, and 140 m along the ditch, measured from the wetland edge (Figure 4.2 inset) at the midpoint
in the water column.
Figure 4.2: Location of the ditches (DT) and spills (SP) sampled, and the extent of open water (includes wetlands
and lakes) determined by Ducks Unlimited Canada from 2005 air photos (courtesy of Lyle Boychuk). Inset: Air
photo of LR3 wetland, showing the sampling sites in the wetland and along the newly constructed drainage ditch
(DR). The location of the water level recorder (PT2X) is indicated by a cross.
4.3 Natural spills and artificial ditches sampling strategy
Seven artificial drainage ditches (hereafter, ditches, DT) and five natural connections
(hereafter, spills, SP) (Table 4.2) were selected for study. The ditches and spills drained wetlands
that did not have any surface water inflows, with the exception of DT3, which had receives
inflow from DT1. However, at the time of sampling, DT1 was not flowing because it was frozen
and snow covered. Due to the variation in ditch and spill length, locations for water sampling
along the ditch varied; they are reported in Table 3.2. Water samples were collected 10-18 April
2009 from the thalweg of the connection, starting at the most downstream sample location.
Samples were obtained from the midpoint of the water column. Manual flow gauging was
carried out at the time of water quality sampling using a Marsh-McBirney Flo-mate 2000
velocity meter and wading rod. Velocity at 60% depth was measured at 20 – 40 cm intervals, so
that no more than 20% of the total stream discharge was measured at each point. Ditch or spill
discharge was then calculated as the sum of the product of velocity and average stream depth at
each sample interval.
15
Table 4.2: Photographs of artificial ditches (DT) and natural connections (SP) studied that drain wetlands at Smith
Creek watershed and means of physical properties measured along the connections: discharge (Q), velocity (v),
depth (d), and width (w). Sample locations were measured from the wetland edge along the connection.
Sample locations (m):
Sample locations (m):
0, 30, 60, 90, 120, 150
0, 25, 50, 75, 100
Q: 0.06 cms
Q: 0.002 cms
v: 0.04 m/s
v: 0.009 m/s
d: 0.37 m
d: 0.07 m
w: 3.05 m
w: 2.99 m
Sample locations (m):
Sample locations (m):
0, 25, 50, 75, 100
0, 25, 50, 75, 100, 125
Q: 0.02 cms
Q: 0.16 cms
v: 0.02 m/s
v: 0.11 m/s
d: 0.33 cm
d: 0.27 m
w: 4.45 m
w: 5.59 m
Sample locations (m):
Sample locations (m):
0, 25, 50, 75, 125, 175
0, 25, 50, 75, 100, 125
Q: 0.02 cms
Q: 0.06 cms
v: 0.01m/s
v: 0.06 m/s
d: 0.22 m
d: 0.32 m
w: 11.0 m
w: 4.01 m
Sample locations (m):
Sample locations (m):
0, 50, 100, 150, 200, 250
0, 5, 10
Q: 0.02 cms
Q: 0.07 cms
v: 0.05 m/s
v: 0.01 m/s
d: 0.16 m
d: 0.28 m
W: 3.05 m
w: 22.3 m
Sample locations (m):
Sample locations (m):
0, 15, 30, 45
0, 8, 15
Q: 0.02 cms
Q: 0.006 cms
v: 0.005 m/s
v: 0.02 m/s
d: 0.16 m
d: 0.07 m
W: 27.1 m
w: 6.40 m
Sample locations (m):
Sample locations (m):
0, 9, 18, 36
0, 10, 20
Q: 0.002 cms
Q: 0.001 cms
v: 0.001 m/s
v: 0.001 m/s
d: 0.18 m
d: 0.10 m
W: 8.49 m
w: 4.57 m
4.4 Stream water quality sampling
Stream water samples were collected at three locations along Smith Creek in 2008, and
four locations in 2009. One sample site was the watershed outlet (SC3), and the other sample
sites were three sub-basins of Smith Creek (SC4, SC5, TV1; Figure 4.3) that represent areas of
the watershed with differing degrees of wetland drainage. Wetland distribution across the
watershed was historically (1958) even (Fang et al., 2010). Although sampling sites were chosen
to reflect differing proportions of wetland loss between 1958 and 2007-2008, quantification of
wetland loss was completed after site selection through the land cover classification of Fang et
al. (2010). Wetland losses in the three subbasins sampled and resulting drainage classes used in
this study are reported in Table 4.3. Stream water quality measurements were taken daily during
peak flow (spring runoff), weekly during low flows, and following heavy summer rains (2008
16
only). Stream discharge at the subbasin sampling sites was estimated at the time of water quality
sampling (2009 only) using a Marsh McBirney flow meter. Streamflow at the watershed outlet
(Figure 4.3) was obtained from Water Survey of Canada gauge 05ME007. The water quality
sampling station representing the watershed outlet was located just upstream of the town of
Langenburg’s drainage (SC3; Figure 4.3) as preliminary chemical analyses showed that it
contributes high fecal bacteria and nutrients to Smith Creek.
Figure 4.3: Location of stream sampling sites and the Environment Canada hydrometric gauge along Smith Creek.
Sub-basins were identified by Fang et al. (2010).
Table 4.3: Wetland loss (%) and corresponding sub-basin drainage classification used in this study. The wetland loss
data are derived from wetland areal extent changes between 1958 (Ducks Unlimited Canada data) and supervised
land use classification data from 2007 and 2008 SPOT5 images (Fang et al., 2010).
Sample Site ID
SC5
SC4
TV1
Wetland Loss 1958-2007/8 in
Sub-Basin (%)
23
57
64
Drainage Class
Low
Medium High
High
4.5 Analysis of water quality
Water samples for bacteria, i.e. Escherichia coli (E. coli) and total coliforms (T. coli),
were only collected for the wetland drainage experiment and from the stream due to challenges
associated with transport times to the laboratory. They were collected in 100 mL sterile bottles,
preserved with Na2S2O3, and submitted to Saskatchewan Research Council Analytical
Laboratories, Saskatoon, SK (SRC) within 24 hours. They were analyzed using the chromogenic
substrate method (Standard Methods part 9223).
Water samples for ion and nutrient analyses were collected in pre-rinsed 1 L polyethylene
or two 350 mL glass bottles, kept on ice during the day, and then split into aliquots.
17
Temperature-compensated specific conductance (SC) and pH were measured in the laboratory on
the day of sampling using Hach sension156 and Orion 3-Star hand-held meters, respectively.
Two sample aliquots were preserved by the addition of HNO3 and H2SO4, respectively, followed
by refrigeration at 4oC. A third aliquot was filtered through a 40 μm Whatman GF/C glass
microfiber filter then frozen. The HNO3 preserved sub-sample was analyzed for total phosphorus
(TP) concentration at SRC within three days of sampling by inductively coupled plasma atomic
emission spectroscopy (Standard Methods part 3120). Samples preserved with H2SO4 were only
collected for the 2008 wetland drainage experiment and analyzed for total Kjeldahl nitrogen
(TKN) at SRC within three days of sampling by digestion and subsequent ammonia analysis
(EPA 351).
Filtered samples were analyzed for a variety of chemical parameters. Total dissolved
nitrogen (TDN) and dissolved organic carbon (DOC) for samples collected in 2009 were
analyzed at the University of Saskatchewan on a Shimadzu TNM-1. DOC samples from 2008
were analyzed at SRC by UV persulfate digestion and non-dispersive IR detection, (Standard
Methods part 5310C). Ortho-phosphate (orthoP) was also analyzed at SRC colorimetrically
(Standard Methods part 4500). A Westco SmartChem Discrete Analyzer (SmartChem 200,
Method 375-100E-1) was used for analysis of ammonium (NH4) and nitrate plus nitrite (reported
as NO3). Major ions (Cl, HCO3, SO4, Na, K, Mg, and Ca) were analyzed by ion chromatography
with a Dionex Model ICS-2000 using potassium hydroxide and methanesulfonic acid EluGen for
anion and cation analysis, respectively, at the University of Saskatchewan. Carbonate
concentrations were assumed to be negligible due to high sample pH and its absence in pilot
tests. One duplicate and one blank sample are analyzed for every fifteen samples collected for
QA/QC. Note that charges have been left off ions throughout this report to enhance readability.
4.6 Ecosystem health sampling sites and frequency
Benthic macroinvertebrate-based biomonitoring approaches are best applied in two time
windows through the course of a year – early spring shortly after ice-off, and fall shortly before
ice-on (Hilsenhoff 1988). Because many Smith Creek streams cease to flow by midsummer,
benthic macroinvertebrates were sampled in late April and early May in 2008 and 2009. Our
2008 study of the Smith Creek provided sampling of four sites on May 8, 2008 (Figure 4.4a;
Table 4.4). However, in 2009 we sampled the stream sites on two occasions early in the spring;
once on April 16, 2009 (8 sites) as the freshet commenced, and later on May 13, 2009 as water
levels declined (Figure 4.4b; Table 4.4). Our first sampling event in 2009 involved collecting
macroinvertebrates from eight sites comprised of four Ditch sites (DT 1, DT 2, DT 3, and DT 4)
and four stream sites (SC 3, SC 4, SC 5, and TV 1). Ditch site DT 5 included in the water quality
and hydrology components of this project was not included here as it was too ephemeral and
observations showed that it lacked macroinvertebrates altogether. In the second sampling for
2009, however, we only repeated sampling at the four stream sites.
18
Table 4.4: Sites, names and coordinates included in the Smith Creek ecosystem health assessment 2008 and 2009.
Stream name Location SWA_Code Smith Creek
Smith Creek
Langenburg Creek
Smith Creek
Smith Creek Tributary
Smith Creek Tributary
Smith Creek Tributary
Smith Creek Tributary
Smith Creek
Smith Creek
Thingvala Creek
Smith Creek
Smith Creek North of Marchwell
Smith Creek East of Langenburg
Langenburg Creek East of Langenburg
Smith Creek at Werle Farm
Ditch at DT3
Ditch at DT1
Ditch at DT4
Ditch at DT2
Smith Creek at SC4
Smith Creek at SC5
Thingvala Creek at TV1
Smith Creek at SC3
SWA_2008_16
SWA_2008_17
SWA_2008_18
SWA_2008_19
SWA_2009_9
SWA_2009_10
SWA_2009_11
SWA_2009_8
SWA_2009_42
SWA_2009_43
SWA_2009_6
SWA_2009_7
UTM13East UTM13North
740829.6
736372.3
734626.6
732536.2
723906.0
723876.0
723870.0
725264.0
726911.7
727319.9
731532.3
736414.4
5636736.0
5638305.1
5637867.9
5640900.0
5645464.0
5645266.0
5645266.0
5652321.0
5648988.6
5648240.1
5641002.9
5638022.5
Figure 4.4: Map of Smith Creek watershed with ecosystem health assessment sampling locations from: a) 2008, and
b) 2009. Note that Smith Creek at Werle Farm is TV1 and Smith Creek East of Langenburg is SC3.
4.7 Macroinvertebrate assemblage sampling methods
At each site, four qualitative 500 μm mesh D-frame net samples were collected along the
length of the reach (defined as 6 times bankful width). Each sample consisted of five, 10 second,
~ 0.3 m2, composite sweeps across the width of the stream; at the left bank, ¼ distance across,
the middle, ¾ distance across, and the right bank. Samples were concentrated on a 500 μm mesh
sieve and preserved in 80 % ethanol in the field. Organisms were returned to the laboratory
where they were sorted from the organic material under 7 X magnification and stored in 80 %
ethanol until they were identified to lowest possible taxon designation. Samples were subsampled where necessary. Final macroinvertebrate abundance was calculated from the
19
subsample fraction and extrapolated to that of the original sample. Samples were required to
produce a minimum of 100 specimens to be included in the data analysis.
Benthic macroinvertebrates were identified to lowest possible designation using keys for
North America (Merritt and Cummins, 1996) and Western Canada (Brooks and Kelton, 1967;
Dosdall and Lemkuhl, 1979; Clifford, 1991; Larson et al., 2000; and Webb, 2002). Voucher
series were deposited in both the Saskatchewan Watershed Authority Invertebrate Voucher
Collection (Saskatoon, Saskatchewan), and the Royal Saskatchewan Museum (Regina,
Saskatchewan).
4.8 Habitat Assessment
Physical characteristics of habitat structure were collected at each site in addition to
benthic macroinvertebrate sampling. Recorded were in-stream substrate composition, vegetation,
algae, and detritus compositions along with notes on surrounding land use and field water quality
measures at the time of sampling. These data and other more detailed GIS information are used
in the test site analysis and reference condition assessment in the future.
The characteristics described in the habitat assessments are summaries and coarse
descriptors of habitat condition at sites, and follow the rapid assessment protocols used by other
jurisdictions in the United States (Barbour et al., 1999). Three main areas were covered in the
site evaluation were substrate (% composition of material, dominant sub-dominant
classification), in-stream habitat (macrophyte, algae, woody debris, and detritus use and
prominence as well as characteristics of embeddedness and riparian cover), and human land use
(riparian land use classification). This information will be used in discerning the condition of the
sample sites when evaluating them in test site analysis in the future.
4.9 Data analyses
4.9.1 Characterization of wetland water quality
A multivariate analysis of water quality was carried out to evaluate the influence of
wetland permanence and surrounding land cover type on resulting wetland water quality. Data
were log transformed to correct for nonnormality and heteroscedasticity (Legendre and
Legendre, 1998). Following preliminary data analysis, the water quality data set was divided into
nutrient (TP, orthoP, TDN, NO3, NH4, DOC, and K) and salinity (SC, pH, Cl, HCO3, SO4, Na,
Mg, K, and Ca) variable sets because the mechanisms controlling nutrient concentrations in the
wetlands have been shown to be different from mechanisms controlling ionic concentrations
(Labaugh, 1987). Mass per volume concentrations (i.e., mg/L) were converted to equivalence
concentrations (i.e., meq/L) for direct comparison of major ions.
A two-way multivariate analysis of variance (MANOVA) was used to test the
relationships among land cover and permanence for the water quality parameters measured. Post
hoc comparison tests (two-way ANOVAs) were used to determine which variables contributed to
the occurrences of significant differences between factors. Significant differences among land
cover and permanence classes were assessed using Tukey's Honest Significant Difference (HSD)
pairwise comparisons test, which is applicable to mildly unbalanced designs (Everitt and
Hothorn, 2006). Statistical analyses were conducted using the R statistical language and
environment (R Core Development Team, 2005). A type I error rate of 0.05 was used in
significance tests unless otherwise stated.
20
4.9.2 Wetland drainage experiment
Wetland volumes were estimated using the full volume-area-depth method (Hayashi and
van der Kamp, 2000) in combination with PT2X water level data. Coefficients required for the
estimates were obtained by Minke et al. (2010) from a digital elevation model derived from total
station survey data. The volume of ice water in the wetland at the time of drainage was estimated
using the density of ice (920 kg/m3) and the average ice thickness (Andres and Van Der Vinne,
2001). Constituent mass in the wetland was estimated using the concentration and the wetland
volume estimate at the time of sampling. Changes in solute concentrations relative to chloride,
which is impacted by evapotranspiration and dilution in the wetland but is biotically
conservative, were used to indicate biotic or geochemical processing of constituents in the
wetland (Duff et al., 2009).
The average concentration along the ditch multiplied by the change in estimated wetland
volume at each sampling point was used to calculate total loads exported from the wetland via
the newly constructed drainage ditch. Loads were normalized along the length of the new ditch
by dividing the load at each sample point along the ditch by the load at the first sampling point
(DR1) in the ditch. A value of one was then subtracted to set the slope intercept to zero. The
statistical significance of the linear relationship with normalized load and distance along ditch
was tested using the linear model (l m) function of the R statistical language and environment (R
Core Development Team, 2005). The slopes of the normalized concentrations were statistically
compared as per Zar (2008) to the slope of biotically conservative Cl, at each sampling instance
to assess whether constituents were added or removed along the length of the ditch. Significantly
different slopes indicate nutrients or salts are either abiotically or biotically removed (or added)
as water travels along the ditch length.
4.9.3 Comparison of ditches and spills
Statistically significant differences in water quality constituents (concentrations and
loads) and physical properties between ditches and spills were assessed with t-tests computed
using SPSS (version 14.0). Data used were log transformed, means measured along each ditch or
spill. Differences between loads at the connection inlet and outlet were also compared using ttests to assess whether transformations of water quality constituents occurred along the length of
the connection. A type I error rate of 0.05 was used unless otherwise stated.
4.9.4 Stream water quality
To account for differences in sub-basin size and water flows, export coefficients (Eij) for
nutrients and salts were calculated by
Eij =
Cij Qij
(1)
Ai
where C and Q are concentration and discharge measured on day j, and A is the area of subbasin
i. This permitted direct statistical comparison of findings among subbasins characterized with
differing wetland drainage. Eij values were compared during the 2009 snowmelt period among
subbasins using a one-way ANOVA and Sidak post-hoc comparisons in SPSS v.14.0.
Additionally, total water and solute exports for spring melt in 2009 were computed by summing
daily values. Solute data from 2008 are not reported because of lack of discharge measurements
21
at all but site SC3 due to equipment failure. However, trends in water quality concentrations in
2008 were similar to those reported for 2009.
4.9.5 Macroinvertebrate analysis
For a complete description of the construction of SWA’s RCA and Test Site Analysis
(TSA) based aquatic ecosystem health tool please refer to the Aquatic Health Feasibility study
(Phillips and McMaster, 2010). A summary of tool construction is provided below.
Reference sites included in this study are a subset of 100 sites collected in 2006 and 2007
as part of the construction of SWA’s aquatic health feasibility study (Phillips and McMaster,
2010). Reference sites were separated from a larger set of sites based on the amount of human
land-use in a buffer area 10 km long and 1 km wide on either side of the stream, upstream of the
site. Because of the predominance of agricultural development in southern Saskatchewan
watersheds, reference condition is not identified as pristine land-cover with the absence of
human activity, but more practically at that condition available that has the ‘least’ amount of
human activity (Stoddard et al., 2006). Specifically, reference site candidacy was restricted to
the following criteria: < 50% cropland, < 5% urban land-use, < 50% pasture, < 80% total land
under human influence (combination of cropland, urban, and pasture), ≤ 2 landfills, oil wells,
bridges or road crossings upstream, and no reservoirs within 10 km. Groups of reference sites
with similar community compositions were identified using cluster analysis, and used to define
what underlying physiochemical conditions structure macroinvertebrate communities in southern
Saskatchewan. In these analyses, it was found that surficial geology composition in the 10 km
buffer upstream of the site, and the site ecoregion membership best explained the composition of
the macroinvertebrate community. From this step, test sites (potentially disturbed sites) were
matched to the appropriate reference group based on similarities in surficial geology and
ecoregion. The test site was then compared with the reference grouping for which it had the
highest probability of membership, determined using discriminate functions analysis. All Smith
Creek Test Sites included in this study were identified as belonging to the Biological Grouping 2,
and thus further analyses comparing the test sites to reference sites were conducted against the
pool of Biological Grouping 2 reference sites (c.f. Phillips and McMaster, 2010).
Next, attributes of the benthic macroinvertebrate assemblage were selected for
comparison between test and reference sites using those adopted as part of the aquatic health
feasibility study (Phillips and McMaster, 2010) and 2010 SWA State of the Watershed Reporting
(SWA 2010). Specifically, the biological endpoints, or metrics, of abundance, taxa richness,
Shannon’s Diversity, and community structure were summarized using Correspondence Analysis
(CA) ordination axes. The measure of taxa richness used was produced as Margalef's species
richness for each sample, and was a measure of the number of species making some allowance
for the number of individuals. Both taxa richness and Shannon’s Diversity were calculated using
PRIMER version 6 (Plymouth Marine Labs, Plymouth, UK; Clarke and Warwick, 2001). We
conducted our CAs using the Biplot add-in for Excel 2009. Separate CA analyses were
conducted for each Smith Creek Test Site with the combination of Reference Sites. The resulting
axis 1 score was used in further test site analysis (Michelle Bowman, personal communication).
Finally, TSA was used to formally evaluate the magnitude of difference between each
test site and the appropriate reference sites (e.g., Bowman and Somers, 2006) based on the
metrics mentioned above. TSA based on individual metrics is equivalent to evaluating whether a
metric value is outside the normal range of variation in the reference sites. A non-central alpha
value (ncP) > 0.95 indicates a site is significantly within reference condition and healthy, a value
22
between 0.95 and 0.05 indicates the site is outside the normal range and is stressed, while a ncP
value ≤ 0.05 indicates that the site is significantly outside the normal range of reference variation
and is thus impaired.
5 Results
5.1 Chemical characterization of wetlands
Nutrient concentrations in the wetlands studied ranged widely (Appendix A; Table 5.1).
TP ranged from 0.02 to 2.8 mg/L. Most wetlands studied were classified as eutrophic, with the
exception of six which were hypereutrophic (i.e. TP > 0.6 mg/L), based on trophic classification
presented in Wetzel (2001). The majority of P was typically in the organic form, with the
exception of six cropped wetlands and one wood wetland that were characterized by greater
proportions of orthoP. TDN in the wetlands ranged from 0.8 to 2.8 mg/L. N was predominantly
present in the organic form with DON making up 96 % on average of TDN. Comparing
DIN:DIP to the Redfield Ratio allows for determination of the limiting nutrient on algae growth
(Rhee and Gotham, 1980; Wetzel, 2001). Eight wetlands were P limited (DIN:DIP>12:1), 46
were N limited (DIN:DIP<7), and 13 were limited by neither N nor P (i.e., DIN:DIN 7-12). DOC
ranged from 19 to 55 mg/L.
Salt concentrations in the wetlands studied were also greatly varied (Appendix B; Table
5.1). The wetlands were neutral to slightly basic (pH of 6.6-8.6) and ranged from fresh to
brackish (SC varied from 57 to 1780 μS/cm). Maximum concentrations of major ions were
HCO3 =3.8 meq/L, SO4=15.5 meq/L, Mg =13.7 meq/L, and Ca =6.8 meq/L. Distinct patterns of
ion dominance groups were apparent (Figure 5.1). With the exception of one wetland (W3), all
29 wetlands with SC > 413 μS/cm were characterized by SO4>HCO3>Cl and Mg>Ca>K>Na. All
but two of these wetlands were classed as crop or grassland and 41% and 34% were classed as
permanent and semi-permanent, respectively. The other anion dominance pattern observed was
HCO3> SO4>Cl, and 50% of these wetlands were also characterized by Mg>Ca>K>Na. The
second most common (21%) cation dominance pattern for these wetlands was Ca>Mg>K>Na.
The two-way MANOVA for the nutrient variable set indicated significant differences
existed among land cover types (p = 0.009) and permanence classes (p = 7x10-5). However, there
was no significant interaction between land cover type and wetland permanence class (p = 0.23).
For the salinity variable set, there was a significant difference among land cover types (p = 5x1010
). However, differences among permanence classes (p = 0.11) and the interaction between land
cover type and wetland permanence class (p = 0.46) were not significant.
23
Figure 5.1: Distribution of cation and anion dominance groups as a function of specific conductivity for the 67
wetlands studied grouped by A) land cover and B) permanence classes. Semiperm is semipermanent wetlands.
Subsequent pairwise comparison tests elucidated differences among land cover types and
permanence classes for the nutrient variable set and land cover types for the salinity variable set
(Table 5.1). There were significant differences among land cover types for TP and K, i.e., crop
wetlands had greater TP and K than wood or grass wetlands. Significant differences among
permanence classes for TP, TDN, and DOC were also found. Permanent wetlands had lower TP
than seasonal and semi-permanent wetlands. In addition, TDN and DOC were higher in seasonal
wetlands compared to semi-permanent and permanent wetlands. SC, Cl, HCO3, SO4, Na, Mg,
and Ca concentrations were significantly lower in the wood wetlands compared to the crop and
grass wetlands.
24
Table 5.1: Mean (standard error) of wetlands within each permanence and land cover class, as well as a summary of
results for two-way ANOVAs. Semiperm is semipermanent wetlands. Differing letter superscripts indicate
significantly different (α = 0.05) Tukey’s pairwise comparisons; * and ** denote a statistically significant difference
at α = 0.05 and α = 0.01, respectively.
Variable
Unit
pH
SC
uS/cm
Cl
meq/L
HCO3
meq/L
SO4
meq/L
Na
meq/L
Ca
meq/L
Mg
meq/L
K
meq/L
TP
mg/L
orthoP
mg/L
TDN
mg/L
NO3
mg/L
NH4
mg/L
DOC
mg/L
DIN:DIP
--
Permanence
Permanence Class
Seasonal Semiperm Permanent
7.18a
7.21a,b
7.52b
(0.09)
(0.09)
(0.10)
446
417
601
(81)
(56)
(102)
7.4
5.0
4.5
(2.0)
(1.1)
(0.7)
101.3
96.0
104.6
(9.0)
(7.4)
(9.5)
99.1
93.6
156.2
(37.9)
(20.8)
(37.5)
14.3
11.5
21.4
(5.8)
(3.6)
(6.1)
28.5
23.4
25.9
(6.5)
(3.1)
(3.7)
28.3
26.7
47.3
(8.1)
(4.9)
(9.8)
21.7
19.9
16.7
(2.1)
(2.7)
(1.1)
0.36a
0.34a
0.11b
(0.08)
(0.14)
(0.02)
0.06
0.09
0.04
(0.02)
(0.03)
(0.01)
1.58a
1.21b
1.12b
(0.10)
(0.07)
(0.05)
0.02
0.04
0.02
(0.00)
(0.01)
(0.01)
0.026
0.024
0.030
(0.003)
(0.003)
(0.003)
38.0a
31.5 b
28.8 b
(2.0)
(1.2)
(1.4)
4.3
6.8
5.6
(1.0)
(2.4)
(1.3)
Two-way ANOVA
F
p
4.26
0.0188*
1.68
0.1952
0.67
0.5158
0.17
0.8423
1.56
0.2187
2.15
0.1260
0.08
0.9206
2.54
0.0880
1.44
0.2464
8.73
0.0005**
0.20
0.8214
11.43
0.0001**
0.16
0.8513
1.02
0.3661
8.96
0.0004**
0.30
0.7422
Land Cover
Land Cover Class
Two-way ANOVA
Crop
Grass
Wood
F
p
7.33c
7.42 c
7.18d
1.96
0.1507
(0.08)
(0.09)
(0.12)
633 c
649 c
194 d
22.86
0.0000**
(104)
(75)
(23)
6.7 c
8.1 c
1.8 d
19.77
0.0000**
(0.8)
(1.9)
(0.2)
115.4 c 114.5 c
72.3 d
10.22
0.0002**
(9.0)
(7.7)
(6.5)
155.5 c 180.9 c
15.9 d
22.07
0.0000**
(39.0)
(36.0)
(6.3)
13.2 c
32.0 c
1.6 d
31.60
0.0000**
(3.8)
(7.0)
(0.5)
33.2 c
31.9 c
12.7 d
13.15
0.0000**
(4.7)
(5.4)
(1.2)
44.1 c
50.2 c
9.4 d
20.72
0.0000**
(10.1)
(8.1)
(2.0)
24.8 c
16.7 d
16.9 d
4.17
0.0203*
(2.7)
(1.2)
(1.4)
0.46 c
0.18 d
0.15 d
4.68
0.0131*
(0.14)
(0.04)
(0.03)
0.10
0.02
0.07
1.98
0.1479
(0.03)
(0.01)
(0.03)
1.40
1.23
1.27
2.62
0.0817
(0.08)
(0.10)
(0.06)
0.03
0.02
0.02
0.54
0.5853
(0.01)
(0.01)
(0.01)
0.026
0.027
0.026
0.56
0.5720
(0.004) (0.003) (0.026)
34.0
32.3
31.7
0.39
0.6822
(2.1)
(1.6)
(1.6)
4.3
6.7
5.3
1.75
0.1833
(1.2)
(2.0)
(1.4)
5.2 Wetland drainage experiment
The water stored in wetland LR3 increased following snowmelt from 776 m3 on October
23, 2007 to 2703 m3 on April 18, 2008. Water volume tended to decrease during rain free periods
and increase following rain events throughout the 2008 open water season (Figure 5.2).
Following the snowmelt period, water volume decreased ~60% throughout spring and early
summer, reaching a minimum on July 7. Frequent and large rain events occurred between July 7
and August 14, which caused water volume in LR3 to increase above that in spring and remain
high until the drainage experiment in November. Based on daily wetland water level fluctuations
(data not shown) and precipitation data, ~6 mm/day of water was estimated to be lost from the
wetland between May 1 and October 22, 2008.
25
Figure 5.2: Daily rainfall and volume in wetland LR3 prior to and during drainage.
Concentration (Figure 5.3) and mass (Figure 5.4) of forms of nitrogen and phosphorus
were highly seasonal variable. TP concentrations in LR3 ranged from 0.22 mg/L to below
analytical detection limits (i.e., 0.01 mg/L). TP mass and concentration were both elevated April
18 and decreased sharply to lows on May 8. During the relatively rain free spring and early
summer, TP concentration increased, peaking July 10. TP mass also increased during the
relatively rain free period, however, it continued to increase during the July and August rain
events whereas TP concentration declined in late July and increased in early August. TP
concentrations and mass stabilized throughout much of August and September and began to
decline in late September. TP mass and concentration spiked October 15 following a 15.8 mm
rain event on October 13-14. OrthoP concentrations in the wetland ranged from 0.15 mg/L to
below analytical detection limits (i.e., 0.01 mg/L). Seasonal variations in orthoP concentration
and mass were similar to those for TP. TP (p = 0.42) and orthoP (p = 0.29) were not significantly
correlated with Cl. The normalized mass data (Figure 5.5) show that TP and orthoP were both
added to the wetland relative to Cl during the growing season.
26
Figure 5.3: Concentration of nitrogen, phosphorus, and bacteria measured in the LR3 wetland prior to the drainage
experiment and in the newly constructed ditch.
27
Figure 5.4: Total mass of nitrogen, phosphorus, and bacteria measured in the LR3 wetland prior to the drainage
experiment and the cumulative mass exported via the newly constructed ditch.
28
Normalized Mass
30
20
TP
10
OrthoP
TKN
5
NO3
NH4
Cl
0
2.5
Normalized Mass
Mg
2.0
Ca
Na
1.5
K
SO4
1.0
HCO3
0.5
DOC
Cl
0.0
May
Jun
Jul
Aug
Sep
2008
Oct
Nov
Dec
Figure 5.5: Mass of various water quality parameters, normalized to May 1, 2008, measured in wetland LR3.
TKN concentrations in the wetland ranged from 3.7 mg/L to 1.1 mg/L. NO3 and NH4
concentrations ranged from below analytical detection limits (i.e., 0.01 mg/L) to 3.6 mg/L and
0.9 mg/L, respectively. Average concentrations of TKN, NO3, and NH4 were respectively 1.9
mg/L, 0.23 mg/L, and 0.06 mg/L. TKN, NO3, and NH4 concentrations were elevated following
snowmelt and masses decreased to lows May 22, May 15 and May 1, respectively. During the
relatively rain free period of April 18 to June 6, TKN concentration increased, peaking July 3.
NH4 and NO3 concentrations were quite variable throughout the study period. At the beginning
of the rain events (July 10) TKN and NO3 decreased substantially while NH4 increased slightly.
Peaks in concentrations were reached July 17 (TKN and NH4) and July 23 (NO3). Minimum
NH4 and NO3 concentrations occurred October 8 and 22, respectively. Mass of TKN and NH4
increased steadily from July 10 until August 21 and August 28, and then tended to decrease until
the start of the drainage experiment. NO3 mass was highly variable during the same time period,
but remained constant and low after September 24. Comparing mass ratios of dissolved inorganic
nitrogen (NO3 + NH4) to dissolved inorganic phosphorus (PO4) (i.e. DIN:DIP) to the Redfield
Ratio indicates that the limiting nutrient in the wetland was P in spring (Figure 5.6; Wetzel,
2001). After June 5, the limiting nutrient in the wetland was predominantly N for the remainder
of the study. TKN was positively correlated with Cl (r = 0.56, p = 1x10-3), however, NH4 (p =
0.72) and NO3 (p = 0.64) were not correlated with Cl. The normalized mass data (Figure 5.5)
show that over the course of the study period, NH4 was added to the wetland, NO3 was removed,
and TKN was neither greatly added nor removed relative to Cl.
29
600
DIN:DIP Ratio
500
50
40
30
20
12:1
7:1
10
0
May
Jun
Jul
Aug
Sep
Oct
Nov
2008
Figure 5.6: Mass ratio of DIN:DIP (NO3+NH4:PO4) measured in the LR3 wetland prior to its drainage. Values
below seven indicate N limiting conditions and values above 12 indicate P limiting conditions (Wetzel, 2001).
Variations in density (Figure 5.3) and loads (Figure 5.4) of both T. coli and E. coli were
very similar throughout 2008, and as such only trends in loads are summarized. Following the
snowmelt period, loads of T. coli and E. coli in LR3 decreased until May 28, then increased until
June 26 and June 19, respectively. Minimums were reached July 10. Loads of T. coli and E. coli
then increased during the mid summer rain events. T. coli and E. coli loads then generally
decreased until the start of the drainage experiment. A secondary peak in E. coli and T. coli
occurred October 8. Neither T. coli (p = 0.12) nor E. coli (p = 0.50) were significantly correlated
with Cl.
Seasonal variations in major ion and DOC concentrations were similar throughout
summer 2008, with the exception of HCO3 (Figure 5.7). DOC (r = 0.88, p = 3x10-11) and all
major ions (r = 0.93-0.98, p = 3x10-25-7x10-14), with the exception of HCO3 (p = 0.32), were
significantly correlated with Cl. Following snowmelt and during the relatively rain free spring
and early summer, ion and DOC concentrations generally increased, peaking around July 10. Ion
concentrations then reached minimums on July 30 during rain events and then increased until the
start of the drainage experiment. In contrast, HCO3 concentration did not noticeably increase
between the snowmelt and the end of July. Steady increases in HCO3 concentrations were
observed between August and November. Trends in major ion and DOC mass (Figure 5.8)
differed from trends in concentration. DOC and major ion mass decreased in the wetland during
the relatively rain free spring and early summer until July 10 then increased on average by a
factor of 2.4 and peaked ~August 28. Major ion and DOC mass mostly decreased in September
and increased in October. The ion dominance pattern in the LR3 wetland was SO4>HCO3>Cl
and Mg>Ca>Na>K, which remained constant throughout the study period in 2008. Seasonal
variation in pH (Figure 5.7) was similar to that observed for concentrations of major ions.
However, pH peaked at 9.5 on July 3 and reached a minimum value of 7.4 August 14. The
normalized mass data (Figure 5.5) show that DOC, HCO3, and Ca were added to the wetland
relative to Cl.
At the time of the drainage experiment, the LR3 wetland was covered by ~8 cm of ice.
As a result, only 81% of the water was in liquid form. The wetland volume decreased rapidly
when the drainage ditch was completed (Figure 5.2). After 4 hrs of drainage, 30% of the water
(total) had exited the wetland via the ditch. Concentrations of N, P, bacteria, and major ions were
greater in the drainage ditch than those measured in the wetland at the start of drainage (Figures
5.3 and 5.4). DOC concentrations in the ditch were less than that measured in the wetland at the
start of drainage, not including the final sampling instance, 23 hr after the start of drainage. The
30
pH was consistently lower in the newly constructed drainage ditch than in the wetland at the start
of drainage (Figure 5.7).
Figure 5.7: pH and concentration of DOC and major ions measured in the LR3 wetland prior to the drainage
experiment and in the newly constructed ditch.
31
Figure 5.8: Total mass of DOC and major ions measured in wetland LR3 prior to the drainage experiment and the
cumulative mass exported via the newly constructed ditch.
Over the course of the drainage experiment, concentrations of N, P, and DOC generally
increased in the drainage ditch. However, concentrations of TP, NO3, and NH4 were greater 1 hr
after the start of drainage compared to concentrations measured 4 and 6 hr after the start of
drainage. Total mass exported from LR3 for each constituent during the experiment is
summarized in Figures 5.4 and 5.8. The total masses of TP, orthoP, NO3, and NH4, and the total
number of E. coli and T. coli colonies exported via the drainage ditch exceeded those estimated
in the wetland at the start of the drainage experiment by a factor of 1.2, 2.1, 19.0, 4.3, 19.4, and
18.9, respectively. Total masses of TKN, DOC, and major ions exported via the drainage ditch
were less than those estimated in the wetland at the start of the drainage experiment by factors
ranging from 0.4 to 0.6.
Results for solutes that were significantly correlated (α = 0.05) with distance along the
drainage ditch and that also had at least marginally significantly different (α = 0.10) slopes than
32
Cl are shown in Figure 5.9. Slopes that differ from Cl suggest that the nutrient or ion experienced
biotic processing, sorption or release along the ditch length. Slopes of DOC (1 hr), orthoP (1 hr,
4 hr, and 23 hr), T. coli (4 hr), HCO3 (6 hr), and NH4 (6 hr and 23 hr) were less steep than the Cl
slope. Slopes of HCO3 (4 hr), T. coli (23 hr), and NO3 (23 hr) were steeper than the Cl slope.
Normalized Cl concentrations were not significantly (α = 0.05) correlated with distance 4 hr and
6 hr after the start of drainage.
Figure 5.9: Slopes of normalized concentrations measured along the newly constructed ditch draining the LR3
wetland at: a) 1 hr, b) 4 hr, c) 6 hr, and d) 23 hr after the start of the drainage experiment. Concentrations were
normalized by dividing the concentration at each sample point along the ditch by the concentration at the first
sampling point (DR1) in the ditch. A value of one was then subtracted to set the intercept to zero. Only constituent
slopes that were at least marginally different (α = 0.10) from the chloride slope and that had significant (α = 0.05)
linear relationships between normalized concentration and distance are shown. P values indicate the level of
significance for slopes differing from chloride, suggesting that a portion of the variability is due to biotic processes
or sorption.
33
5.3 Comparison of ditches and spills
Ditches were on average 71 % longer (p = 1x10-4), 12 % narrower (p = 0.035), and
tended to have 33% higher flow velocities (p = 0.016) than spills (Figure 5.10). Constituent
concentrations also differed between ditches and spills (Figure 5.11). Specifically, TDN (p =
0.003), DOC (p = 0.007), HCO3 (p = 0.023), K (p = 0.001), and Ca (p = 0.010) concentrations
were greater in ditches than spills. NO3 (p = 0.058) and NH4 (p = 0.055) concentrations tended to
be higher in ditches than spills. Loads of TDN (p = 0.038), NO3 (p = 0.034), and K (p = 0.048)
were significantly greater in ditches than spills (Figure 5.12). In contrast, loads and
concentrations of TP and orthoP were not significantly different between ditches and spills. Ttest comparisons of inlet and outlet loads showed that solute loads did not change along the
length of ditches or spills (0.29 < p < 0.97), with the exception of orthoP which tended to be
greater (p = 0.058) at spill outlets than inlets.
Figure 5.10: Mean and standard error of ditch and spill physical properties where * and ** denote statistically
significant differences at α = 0.05 and 0.01, respectively.
Figure 5.11: Mean and standard error of A) nutrients and DOC, and B) major ions, SC, and pH in ditches and spills
where * and ** denote statistically significant differences at α = 0.05 and 0.01, respectively.
34
Figure 5.12: Mean and standard error of ditch and spill nutrient and DOC loads where *denotes a statistically
significant difference at α = 0.05.
5.4 Stream water quality
Typical of the northern prairies, the 2009 snowmelt period was short, lasting 28 days.
Thaw (and subsequent flow) at the low drainage subbasin was delayed by about a week
compared to that at the Smith Creek outlet and its other subbasins. Maximum stream discharge
and concentrations of the water quality parameters measured, with the exception of SC and E.
coli, occurred during onset of the spring freshet (Figure 5.13). Ratios of DIN (NO3 + NH4) to
DIP (PO4) measured near the Smith Creek outlet indicate N limiting conditions on nearly all
sampling dates in 2009 (Figure 5.14). During the recession limb of the hydrograph, the dominant
anion switched from HCO3 to SO4 (Figure 5.15), coincident with an increase in SC to >520
mS/cm (Figure 5.13). This switch was observed in each of the studied subbasins.
There was a trend toward higher discharge, and nutrient and salt concentrations at the
subbasins with medium high and high wetland drainage compared to the subbasin with low
wetland drainage (Figure 5.13). After normalizing concentration data for variations in
streamflows and basin areas, differences in export coefficients among subbasins with variable
wetland drainage were found for the 2009 freshet (Table 5.2). Subbasins characterized by
medium high and high wetland drainage had significantly higher mean export coefficients of TP,
orthoP, HCO3, SO4, Mg, and Ca than the subbasin with low wetland drainage. Because of its
large size, the subbasin characterized by medium high wetland drainage had the highest total
export of nutrient, salts and bacteria (Figure 5.16). However, total export of these water quality
constituents at Smith Creek outlet, which is located just upstream of the town drainage
(Langenburg Creek), was higher than the sum of the three studied subbasins.
35
1.0
0.5
45
Low Drainage
1.4
1.2
1.0
0.8
0.6
0.4
0.2
0.0
10.00
8.00
6.00
4.00
2.00
0.50
15
10
5
2.00
1400
1200
1000
800
600
400
200
0
SC (μS/cm)
NH4 (mg/L)
90
80
20
0
1.00
0.50
0.10
0.00
Jun
30
0.00
0.05
May
35
20
1.50
Apr
40
25
0.25
0.0
OrthoP (mg/L)
50
Smith Creek Outlet
High Drainage
Mod. High Drainage
E. coli (MPN/100 mL)
1.5
NO3 (mg/L)
TP (mg/L)
2.0
14
12
10
8
6
4
2
0
DOC (mg/L)
TDN (mg/L)
Discharge (cms)
7
6
5
4
3
2
1
0
Apr
May
Jun
Apr
May
Jun
2009
2009
2009
Figure 5.13: Discharge and water quality concentrations during the 2009 freshet at sites characterized by differing watershed drainage: high (TV1), medium high
(SC4), and low (SC5). No data are presented prior to April 8, 2009 due to ice at the gauge station.
36
18
16
DIN:DIP
14
12:1
12
10
7:1
8
6
4
2
0
Apr
May
Jun
2009
Figure 5.14: Ratios of DIN (NO3 + NH4) to DIP (PO4) measured near the Smith Creek watershed outlet (SC3)
during 2008 and 2009. Values below seven indicate N limiting conditions while those above 12 indicate P limiting
conditions (Wetzel, 2001).
37
Basin Outlet
High Wetland Drainage
Medium High Wetland Drainage
Low Wetland Drainage
Figure 5.15: Piper plots for: basin outlet (SC3) and sub-basins characterized by high (TV1), medium high (SC4),
and low (SC5) wetland drainage. Blue and red circles indicate samples obtained during the snowmelt period (March
23 – April 19, 2009) and following snowmelt (April 25 – May 26, 2009), respectively.
38
Table 5.2: Mean (standard error) stream export coefficients for the 2009 freshet by site. Differing superscripts
indicate significant difference at α = 0.05, assessed using a one-way ANOVA and the Holm-Sidak method for
multiple pair-wise comparisons.
Variable
Units
Wetland Drainage Intensity ANOVA results
Low Med. High High
F
P
Q
m3/day/km2 392a
1212b
1293b 5.24
0.011
(151)
(154)
(208)
TP
kg/day/km2 0.06a
0.60b
0.55b
5.81
0.008
(0.02)
(0.07)
(0.13)
OrthoP
kg/day/km2 0.01a
0.45b
0.34b
6.94
0.003
(0.01)
(0.06)
(0.09)
DOC
kg/day/km2 14.1a
41.2a,b
49.6b
4.59
0.019
(5.53)
(4.83)
(9.08)
TDN
kg/day/km2 0.93a
2.31a,b
6.02b
5.49
0.010
(0.42)
(0.27)
(1.45)
NO3
kg/day/km2 0.29a
0.63a
2.34a
3.37
0.048
(0.19)
(0.1)
(0.79)
NH4
kg/day/km2
0.1a
0.1a
1.1b
6.38
0.005
(0.06)
(0.01)
(0.31)
Cl
kg/day/km2
1.7a
8.5a
8.0a
3.54
0.042
(0.4)
(0.7)
(2.2)
HCO3
kg/day/km2 34.0a
123.7b
120.9b 4.28
0.024
(11.2)
(17.3)
(24.2)
SO4
kg/day/km2 26.8a
95.0b
82.1b
3.92
0.031
(7.1)
(11.2)
(18.6)
Na
kg/day/km2
3.0a
10.4b
6.3a,b
6.56
0.004
(1.1)
(1.3)
(1.3)
K
kg/day/km2
6.2a
24.7a,b
29.3b
4.98
0.014
(2)
(3.5)
(5.7)
Mg
kg/day/km2
5.4a
23.1b
21.7b
6.43
0.005
(1.8)
(2.7)
(3.8)
Ca
kg/day/km2
9.5a
36.0b
40.4b
5.36
0.010
(3.5)
(3.4)
(7.8)
39
18
10000
300
a)
1000
b)
16
250
800
12
200
10
150
8
6
100
Ions (kg x1000)
14
DOC (kg x1000)
4000
Nutrients (kg x1000)
6000
3
Water (m x1000)
8000
4
2000
600
400
200
50
2
0
0
0
Q
TP
orthoP TDN
NO3
NH4
0
DOC
Cl
Na
K
Mg
Ca
2.0e+14
2.5e+11
c)
2.0e+11
High
Medium High
Low
Unassessed (SC3 subbasin)
1.5e+14
1.5e+11
1.0e+14
1.0e+11
T. coli (MPN)
E. coli (MPN)
HCO3 SO4
5.0e+13
5.0e+10
0.0
0.0
E.coli
T.coli
Figure 5.16: Total exports of: a) water (Q), nitrogen (TDN, NO3, and NH4) and phosphorus (TP and orthoP); b)
major ions; and c) bacteria (E. coli and T. coli) during the 2009 freshet (3 April - 26 May) at sub-basins
characterized by low, medium high, and high wetland drainage. Exports for the SC3 subbasin (c.f. Figure 4.3) were
calculated by subtraction of exports for the three subbasins from that for the whole watershed. Total height of the
bars indicates export for the whole Smith Creek watershed.
5.5 Description of macroinvertebrate assemblages
Acquired were 17 samples representing eight sites in 2008, and 15 samples representing
eight sites across two time periods in 2009. Because we were unable to find suitable sampling
sites through much of the upstream Smith Creek watershed, we focused sampling efforts in 2008
on replication at the site level. Specifically, six samples were produced from the Smith Creek at
Marchwell (SWA_2008_16) site, eight samples were produced from the Smith Creek East of
Langenburg site (SWA_2008_17), three from the Langenburg Creek East of Langenburg Site
(SWA_2008_18), and a single sample from the Smith Creek at Werle Farm Site
(SWA_2008_19). Sample number 3 from the Langenburg Creek East of Langenburg and
samples 2, 3, and 4 from the Smith Creek at Werle Farm all contained too few
macroinvertebrates for further analysis.
Based on the consistency of the results at the Marchwell and Langenburg sites
(SWA_2008_16 and SWA_2008_17 respectively, Table 5.3) in 2008, and constraints of
budgeted sorting and identification costs, only a single sample from each of the sites monitored
was analyzed in 2009. However, the remaining three samples from each of the sites have been
retained for future analysis.
Overall, 27,985 benthic macroinvertebrates were collected, representing 106 unique taxa
(Appendix C). This study was dominated by the fairy shrimp (Anostraca), the dipteran midge
larvae (Chironomidae), freshwater worms (Oligochaeta), nematodes (Nematoda), one genera of
40
mayfly (Caenis) and blackfly larvae (Simuliidae). There were no surprises or occurrences of
unexpected or rare taxa based on previous understanding of prairie streams in the area (see
Phillips et al., 2008).
The 2008 results indicate that the Smith Creek at Marchwell is a healthy site, and not
significantly different from reference conditions (Table 5.3). However, the Smith Creek east of
Langenburg produced three of seven sites which were stressed, the Langenburg Creek was
overall severely impaired, and the Smith Creek at Werle Farm was stressed and close to the
impairment threshold (Table 5.3). A closer look at the nature of the ecosystem health conclusions
based on the individual metrics that make up this TSA identifies that abundance is consistently
within reference condition (Table 5.4).
Table 5.3: Sites sampled for ecosystem health in 2008.
Smith Creek
Smith Creek
Site Code
Watershed Site
SWA_2008_16_1
Smith Creek North of Marchwell
SWA_2008_16_2
Smith Creek North of Marchwell
SWA_2008_16_3
Smith Creek North of Marchwell
SWA_2008_16_4
Smith Creek North of Marchwell
SWA_2008_16_5
Smith Creek North of Marchwell
SWA_2008_16_6
Smith Creek North of Marchwell
SWA_2008_17_1
Smith Creek East of Langenburg
SWA_2008_17_2
Smith Creek East of Langenburg
SWA_2008_17_3
Smith Creek East of Langenburg
SWA_2008_17_4
Smith Creek East of Langenburg
SWA_2008_17_6
Smith Creek East of Langenburg
SWA_2008_17_7
Smith Creek East of Langenburg
SWA_2008_17_8
Smith Creek East of Langenburg
SWA_2008_18_1
Langenburg Creek East of Langenburg
SWA_2008_18_2
Langenburg Creek East of Langenburg
SWA_2008_18_4
Langenburg Creek East of Langenburg
SWA_2008_19_1
Smith Creek at Werle Farm
41
TSA Results
D
ncP
1.86
1.00
1.50
1.00
0.89
1.00
1.28
1.00
0.87
1.00
1.19
1.00
2.39
0.99
1.09
1.00
3.18
0.62
2.21
1.00
4.03
0.12
2.76
0.89
0.93
1.00
7.75
0.00
33.02
0.00
4.09
0.11
4.17
0.09
Overall
Health
Healthy
Healthy
Healthy
Healthy
Healthy
Healthy
Healthy
Healthy
Stressed
Healthy
Stressed
Stressed
Healthy
Impaired
Impaired
Stressed
Stressed
Table 5.4: Metric values and their probability of difference from reference (Metric Significance)
Metric Values
SWA
Taxa
Shannon's
CA
Site
Abundance
Richness
Diversity
Axis 1
SWA_2008_16_1
453
4.42
2.56
0.87
SWA_2008_16_2
235
4.21
2.46
1.02
SWA_2008_16_3
120
3.55
2.24
1.11
SWA_2008_16_4
396
4.01
2.36
0.98
SWA_2008_16_7
58
3.20
2.37
1.15
SWA_2008_16_8
365
3.73
2.08
0.85
SWA_2008_17_1
1295
4.88
2.15
0.80
SWA_2008_17_2
142
3.63
2.01
1.02
SWA_2008_17_3
844
4.75
2.28
0.38
SWA_2008_17_4
483
4.69
2.25
0.79
SWA_2008_17_6
635
4.49
2.28
0.15
SWA_2008_17_7
311
3.14
2.32
0.49
SWA_2008_17_8
496
2.74
1.78
0.75
SWA_2008_18_1
634
2.33
1.54
2.57
SWA_2008_18_2
732
1.21
0.28
-6.63
SWA_2008_18_4
1634
1.89
1.48
1.64
SWA_2008_19_1
28
0.00
0.00
0.87
SWA_2009_10_1
196
1.14
0.84
3.65
SWA_2009_11_1
224
1.29
0.54
0.88
SWA_2009_4_1_May
1102
1.86
1.12
-0.43
SWA_2009_4_3_April
1068
4.45
1.75
0.85
SWA_2009_5_1_April
192
3.23
2.22
1.00
SWA_2009_5_1_May
337
6.36
2.76
0.82
SWA_2009_6_1_May
523
2.56
1.21
0.93
SWA_2009_6_2_April
740
2.88
1.39
0.90
SWA_2009_7_1_May
424
3.31
1.66
0.91
SWA_2009_7_2_April
546
3.12
1.54
0.89
SWA_2009_8_1
70
0.71
0.39
0.86
SWA_2009_9_1
576
3.46
2.58
0.82
42
Abundance
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
Metric Significance
Taxa
Shannon's
Richness
Diversity
0.85
1.00
0.97
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
0.30
1.00
1.00
1.00
0.45
1.00
0.52
1.00
0.77
1.00
1.00
1.00
1.00
1.00
1.00
1.00
0.82
0.01
1.00
1.00
0.01
0.00
0.74
0.43
0.89
0.05
1.00
0.95
0.82
1.00
1.00
1.00
0.00
0.92
1.00
0.99
1.00
1.00
1.00
1.00
1.00
1.00
0.24
0.02
1.00
1.00
CA
Axis 1
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
0.89
1.00
1.00
0.00
0.00
0.86
1.00
1.00
1.00
0.02
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
Early 2009 spring sampling results indicate that all stream sites in the Smith Creek were
healthy regardless of drainage stress (Table 5.5). However, May samples show that SC 4 had
become impaired and SC 5 had become stressed (Table 5.6). Ditches in the 2009 study ranged
from healthy at DT 3, stressed in DT 4, DT 2 and DT 1 (Table 5.6).
Table 5.5: Test Site Analysis results for sites from April 13, 2009 in the Smith Creek Watershed. D indicated the
magnitude of effect or difference from reference, while ncP is the measure of significance.
Smith Creek
Drainage
TSA Results
Overall
Watershed Site
Stress
D
ncP
Health
Healthy
Smith Creek at SC4
Medium High Drainage
2.46
0.98
Smith Creek at SC5
Low drainage
0.66
1.00
Healthy
Healthy
Thingvala Creek at TV1
High Drainage
1.51
1.00
Smith Creek at SC3
Very High Drainage
1.23
1.00
Healthy
Healthy
Ditch at DT3
Ditch
1.76
1.00
Stressed
Ditch at DT1
Ditch
3.00
0.75
Stressed
Ditch at DT4
Ditch
2.99
0.75
Stressed
Ditch at DT2
Ditch
3.30
0.52
Table 5.6: Test Site Analysis results for sites from April 13, 2009 in the Smith Creek Watershed. D indicated the
magnitude of effect or difference from reference, while ncP is the measure of significance.
Smith Creek
Watershed Site
Smith Creek at SC4
Smith Creek at SC5
Thingvala Creek at TV1
Smith Creek at SC3
Drainage
Stress
Medium High Drainage
Low drainage
High Drainage
Very High Drainage
TSA Results
D
ncP
5.54
0.00
3.88
0.18
1.82
1.00
1.28
1.00
Overall
Health
Impaired
Stressed
Healthy
Healthy
6 Discussion
6.1 Wetland water quality characterization
Researchers have previously used ion dominance patterns and SC (as a proxy for net
groundwater seepage rates) as indicators of wetland permanence (Sloan, 1972; Millar, 1976;
Driver and Peden, 1979). Although exchanges between the wetland and deep groundwater have
only a minimal influence on the water balance, their effect on salinity can be important as the
direction of flow determines whether salts accumulate in the wetland due to upward flow or are
leached out of the wetland by outward flows (Hayashi et al., 1998b; van der Kamp and Hayashi,
2009). SC has thus been used as an indicator of relative position of wetlands along regional
groundwater flow paths (LaBaugh and Swanson, 2004). However, results presented herein show
that neither SC nor ion dominance can be used to distinguish among wetland permanence classes
at Smith Creek watershed. Although five of the six highest SC measurements were obtained
from permanent wetlands, there was no significant difference among permanence classes for SC
or solute concentrations. The high within permanence class variation likely occurred because
localized shallow groundwater flow systems dominate the groundwater hydrology of prairie
43
potholes (Hayashi et al. 1998a; van der Kamp et al., 2008). Thus, linkages between SC and
permanence may only be appropriate for comparisons made at the regional scale, and the focus
of this study was at a smaller spatial scale. Furthermore, solute concentration and SC were likely
diluted during the sampling period by snowmelt runoff, which would mask differences between
permanence classes. Differences among permanence classes may however become more
pronounced later in the season due to evapoconcentration (Rózkowska and Rózkowski, 1969;
LaBaugh et al., 1987). Other studies have mostly conducted their sampling campaigns midsummer (e.g. Arts et al., 2000); however, wetlands included in this study were not sampled later
in the season because seasonal wetlands typically become dry. The lack of difference among
permanence classes may also be attributed to the fact that classes are not necessarily strictly
distinct and static. Instead, potholes within each permanence class can show sharp differences in
time of inundation and in water depth between periods of drought and deluge (Johnson et al.,
2004), conditions common in the prairies.
The wetlands sampled grouped distinctly into low SC and HCO3 dominated groups,
largely typical of the wood wetlands, and relatively high SC and SO4 dominated groups, largely
typical of grass and crop wetlands. An association between SC and anion dominance has been
previously observed in prairie lakes and wetlands (Rawson, 1944; Rózkowska and Rózkowski,
1969; Barica, 1975; LaBaugh et al., 1987, Swanson et al., 1988). As solutes become increasingly
concentrated by evaporation in closed basins, saturation levels for calcium and magnesium
carbonates are reached first, and then of calcium sulfate, causing the minerals to precipitate in
that order (Holland, 1978). The significantly lower SC and solute concentrations measured in the
wooded areas may be a result of a high order topographic position of these wetlands in the Smith
Creek watershed and thus their likely position near the top of regional groundwater flow paths
(Tóth, 1963; Rózkowski, 1969; Devito et al., 2000).
Topography, however, is not always a good indicator of water table elevations in the
prairies (van der Kamp, pers. comm.). Transects studies by Barica (1978) indicate that closed
basin lake salinity varied considerably regardless of the general topographic slope. Also, wood
wetlands may, in part, have lower Ca, K, and Mg concentrations than grass or crop wetlands
because they uplands are comprised primarily of aspen (Populus tremuloides), which are known
to store these solutes in their standing biomass (Wang et al., 1995).
SC and solute concentrations measured throughout Smith Creek were lower than
maximum values measured at other lakes and wetlands in the prairie pothole region (Rózkowska
and Rózkowski, 1969; Barica, 1975; LaBaugh et al., 1987; Detenbeck et al., 2002; Waiser, 2006)
and were comparable to values observed by Nicholson (1995) in a northern Alberta transition
zone between semi-arid prairie and moister boreal forest. These lower concentrations may be the
result of dilution caused by the relatively higher annual precipitation at Smith Creek watershed
compared to other parts of the prairie pothole region (Millet et al., 2009), which would minimize
the precipitation to evaporation ratio and the effects of evapoconcentration. Additionally, the
mid-summer precipitation events that occurred in the region during the 2008 summer prior to
wetland sampling caused many potholes to fill with dilute rain water and remain relatively full at
the time of freeze-up. SC was also likely low because samples were collected in late spring
shortly after the wetlands have filled with dilute snowmelt runoff and before solutes have
become concentrated by evaporation. Dilute salt concentrations in precipitation have been
measured 220 km southwest of Smith Creek at Bratt’s Lake station, which is part of the
Canadian Air and Precipitation Monitoring Network (CAPMoN, 2007). Although not measured,
it is likely that some of the HCO3 dominated wetlands with SC near 400 μS/cm became SO4
44
dominated later in summer. This inference is based on the work of LaBaugh et al. (1987) and
Detenbeck et al. (2002), which showed some temporal progression in seasonal wetland ion
dominance patterns.
Some of the spatial variation in some wetland nutrients can be attributed to variations in
land cover and permanence classes. The higher concentrations of TP, TDN and DOC in seasonal
wetlands likely results from the pronounced periods of flooding and drying that affect them.
Mineralization of organic matter is enhanced during dry periods, resulting in leaching of
nutrients from standing dead litter and sediment when the next inundation occurs (Bärlocher et
al. 1978; Neill, 1995; Baldwin and Mitchell, 2000; Aldous et al., 2005). In contrast, the lack of
vegetation and continuous flooding of sediments located within the open-water zones of
permanent wetlands and some semi-permanent wetlands would not lead to the same annual
release of nutrients (Brinson et al., 1981). The near continuous flooding of semipermanent and
permanent wetlands suggests that conditions may be sufficiently reduced for denitrification to
occur, also leading to comparatively lower total N concentrations (Neely and Baker, 1989;
Crumpton and Goldsborough, 1998). Seasonal wetlands have been shown to have higher TP
concentrations compared to semi-permanent ones due to leaching from standing dead plants in
the marsh zones surrounding them (LaBaugh and Swanson 2004). Further, seasonal wetlands are
often cropped when they are dry, thus they can have higher N and P concentrations than
semipermanent or permanent wetlands because of direct fertilizer application (Cowardin et al.,
1981).
In addition to differences in wetland nutrients among wetland permanence classes,
differences in some nutrients were found among wetland land cover types. The greater TP and K
concentrations in crop than wood and grass wetlands may have resulted from fertilizer inputs
being transported to wetlands from their cropped watersheds (Hansen et al., 2002; Little et al.,
2007; Tiessen et al., 2010) during snowmelt. The finding that N and DOC did not differ
significantly among land cover types was surprising given that other researchers have shown that
wood litter contains fewer nutrients and DOC than grass litter (Fuller and Anderson, 1993;
Köchy and Wilson, 1997), and that nutrients stay stockpiled in the woody biomass located
throughout wood wetland watersheds for extended periods of time (Wang et al., 1995). The lack
of significant difference in N among land cover types may be attributed to the fact that the
majority of wetlands sampled, similar to shallow prairie lakes, were eutrophic and characterized
by low DIN:DIP ratios and were thus N limited (Barica, 1990; Hall et al., 1999).
The greater TP and K concentrations measured in crop than wood and grass wetlands
may also be the result of varying amounts of surface runoff from different land cover types.
Surface runoff and the nutrients transported with it are likely greatest in cropped areas partly
because of higher overland flows during spring (van der Kamp et al., 2003; Bodhinayake and Si,
2004). The infiltration potential of frozen soils is higher if the soils are dry and have a well
developed macropore structure compared to saturated soils with poor macropore development,
which have very low infiltration potential (Gray et al., 2001). Cultivation reduces macroporosity
and infiltration capacity (Bodhinayake and Si, 2004), which increases infiltration of snowmelt
water and rain, and decreases surface runoff to depressions (van der Kamp and Hayashi, 2009).
Further, snow and rain interception by forest canopies reduce snowpacks and throughfall
(Pomeroy et al., 1997). Consequently runoff in wood areas is likely lower than grass and crop
areas.
It is unlikely that the observed differences in wetland salinity and nutrient concentrations
among land cover types were attributed to differences in soil characteristics. While spatial
45
variation in soil types occur across the Smith Creek watershed (Figure 4.1), the majority of
wetlands resided in soil units characterized by orthic or calcareous soils on lower slopes or
poorly drained depressions. In all, 100%, 46%, and 50% of crop, grass, and wood wetlands,
respectively, were located in soil units with orthic soils in depression, 17% and 36% of grass and
wood wetlands were located in soil units with calcareous soils in depression, and 38% of grass
wetlands and 14% of wood wetlands were located in soil units with carbonated and mixed (orthic
and calcareous) soils in depressions. There was a trend, however, between SC-SO4 dominance
and soil type. Forty one percent of wetlands characterized by high SC and SO4 dominance were
located in soil units characterized by slightly larger proportions of sandy loams (OxWh2,
OxWs2, and Yk12). These soils are expected to have a relatively higher hydraulic conductivity
which means that the wetlands would likely experience greater shallow groundwater and salt
fluxes.
Results also show that the interaction between land cover and permanence classes did not
significantly influence nutrient or salt water quality parameters. This result is not surprising
given that the control mechanisms on water quality parameters attributable to wetland
permanence and land cover type are likely additive. The lack of significant interaction may have
also resulted from the relatively high variability of wetland water quality measured. Other factors
not considered in this study that could also influence prairie wetland water quality include
grazing, cultivation, and tillage practices, varying groundwater fluxes, and the presence of
willow rings surrounding wetlands.
6.2 Factors controlling temporal patterns in wetland water quality
The LR3 wetland effectively trapped N, P, DOC, coliforms, and major ions during runoff
events and exchanged them with the surrounding uplands between events prior to the
construction of the drainage ditch. Intensive temporal measures at the LR3 wetland suggest
hydrological processes are the dominant control on all the studied ionic water quality parameters,
except for HCO3. However, the lack of significant correlations with Cl for most nutrient
variables and differing seasonal dynamics of concentrations, masses and normalized masses
between major ions and DOC compared to N, P, and bacteria suggest there are other processes
important in determining wetland nutrients. This concept of differing control mechanisms for
major ions and nutrients was hypothesized by LaBaugh et al. (1987). Seasonal fluctuations of
ions and DOC appear to be primarily linked to hydrological processes due to the significant
correlations with Cl. Instead, variable seasonal dynamics observed between major hydrological
events and the lack of significant correlations between Cl and TP, orthoP, NH4, NO3, and
bacteria suggests that wetland N, P, and bacteria are linked to both hydrological processes and
biotic/sorption processes. For example, sequences of algae and plant uptake and decay, microbial
processing (i.e. mineralization, nitrification, and denitrification), sedimentation, and waste from
waterfowl and semi-aquatic mammals have been all shown to influence wetland nutrient
availability (Barica, 1974; Neely and Baker, 1989; Neill, 1995; LaBaugh and Swanson, 2004).
Removal mechanisms for bacteria in wetlands include sediment retention and natural die-back
(Hemond and Benoit, 1988; Auer and Niehaus, 1993).
Temporal variations in solute concentrations and masses in the wetland can largely be
attributed to the hydrologically isolated nature of the wetland. The estimate of ~6 mm/day of
water lost from the wetland is within the range of those for other isolated prairie wetlands (1.7 –
7.4 m) via shallow groundwater infiltration and evapotranspiration (Shjeflo, 1968; Millar 1971;
Woo and Roswell, 1993; Hayashi et al. 1998a; Su et al., 2000). Shallow, lateral flows, driven by
46
hydraulic gradients and transpirative demand by upland plants, occur at wetland margins within
the top 5-6 m of till in this landscape where the hydraulic conductivity is relatively high due to
fractures (Hayashi et al., 1998a). These flows transport solutes with them (Hayashi et al. 1998b;
Parsons et al., 2004), explaining the reduction in mass of most water quality constituents in the
wetland between rain events.
Transport with lateral flows can effectively concentrate solutes in uplands (Arndt and
Richardson, 1993; Winter and Rosenberry, 1995). They are then likely to be leached and cycled
back to the wetland during snowmelt and rainfall runoff events. Precipitation events can also
cause the water table beneath the wetland margin to rise above the pond level causing a reversal
in shallow groundwater flow toward the pond (Gerla 1992; Winter and Rosenberry, 1995;
Hayashi et al., 1998b; Parsons et al., 2004), which would also transport solutes back to the
wetland. Daily water level increases in the LR3 wetland exceeded daily precipitation by >5 mm
on 14 days during the study period, indicating that surface and/or subsurface runoff likely
contributed to the increase in wetland water storage. Specifically, runoff contributions to the
wetland likely occurred on at least nine days during the midsummer rain events as well as on
May 28, June 12, June 23, October 6 and October 12. Surface and/or subsurface runoff are thus
likely responsible for the high solute masses and bacteria loads observed during the snowmelt
period and the increases that occurred following the rain events.
Concentrations of DOC and major ions increased during rain free periods likely due to
evapoconcentration and mass decreased likely due to the transport with water out of the wetland
by means of shallow groundwater seepage (Hayashi et al. 1998b; Waiser, 2006), as is described
above. Data from Bratt’s Lake station of the Canadian Air and Precipitation Monitoring Network
(CAPMoN), which is 220 km southwest of the Smith Creek watershed, show precipitation is
dilute with regards to major ions (CAPMoN, 2007). As a result, inputs during the midsummer
rain events probably led to DOC and major ion dilution. It is thus likely that the increases of
DOC and major ions mass measured immediately following rain events were caused primarily
by runoff and shallow groundwater inputs.
Following snowmelt, the wetland had high nutrient concentrations. While transient,
elevated N and P concentrations immediately following snowmelt are important to note because
ditched wetlands drain at this time. Similar to results from Batt et al. (1989) and LaBaugh and
Swanson (2004), these elevated N and P concentrations coincided with periods of use by
breeding waterfowl. They also are coincident with a 2.5-fold increase in wetland volume, relative
to the previous fall. Freshly flooded above ground litter leaches stored N and P to the water
column (Reddy and Patrick, 1975; Davis and van der Valk, 1978; Neill, 1995) and nutrients
accumulated in soils of the wetland periphery are released upon re-wetting (Reddy and Patrick,
1975; Murkin et al., 2000; Aldous et al., 2005). Soils in the region are naturally nutrient rich,
especially in P (Anderson, 1988). In addition, nutrients were likely released from the cropped
areas of the catchment that received fertilizer in fall 2007 and May 2008, as has been shown to
occur in other agricultural systems (Hansen et al., 2002; Little et al., 2007).
The midsummer runoff events led to increased TKN and TP mass in the LR3 wetland.
Prior to the rain events orthoP represented 13% on average of TP and increased to 60%
following the rain events. This increase in the proportion of TP as orthoP may have resulted from
the transport of orthoP with sediment (Neely and Baker, 1989) during overland flows. Although
NH4 and NO3 are predominantly transported with surface runoff and subsurface flows,
respectively (Neely and Baker, 1989), their percentages of TN were similar before and during the
midsummer rains. Average concentrations of NO3 and NH4 in precipitation at the CAPMoN
47
Bratt’s Lake station (2007) are 2.1 mg/L and 0.95 mg/L, respectively. These concentrations are
much greater than those measured in the wetland throughout the study period. Thus wet
deposition in addition to upland runoff and leaching from soil and plants located within the
freshly flooded wetland periphery could have contributed to the increase in nutrient
concentrations and mass loads measured in midsummer. The normalized mass data (Figure 4.4)
suggest that reductions in TP, orthoP, and NH4 mass following the midsummer rain events were
due to biotic uptake and/or biogeochemical reactions rather than hydrological processes.
Furthermore, these data also suggest that NO3 was removed from the wetland relative to Cl,
likely by biotic uptake and denitrification. Although measurements of nitrogen cycling in prairie
wetlands are limited (e.g. Moraghan, 1993), researchers such as Neely and Baker (1989) have
noted that conditions are likely suitable for denitrification to occur: anaerobic conditions and the
presence of a large organic carbon stock.
Temporal variations in HCO3 were likely influenced by carbonate equilibrium
relationships as Heagle et al. (2007) identified carbonate mineral dissolution to be an important
geochemical reaction in a recharge prairie wetland. Normalized mass data showed that
proportions of HCO3 and Ca became elevated relative to Cl. This result suggests that the increase
in these ions was not attributable to water inputs, marked by Cl variations, alone. Sulfate
reduction was also identified as a key geochemical reaction by Heagle et al. (2007). The
proportional mass data did not indicate that SO4 was removed from the wetland differently than
was Cl, meaning sulfate reduction was likely not an important driver of SO4 concentrations at
this site.
A major non-point source of disease causing coliforms and indicator coliforms in
agricultural landscapes is runoff containing animal wastes from pastures or fields fertilized with
manure (Hyland et al., 2003). While the wetland catchment was not fertilized with manure over
the course of the study period, the high coliform densities measured have been observed in
comparable agricultural systems following snowmelt and large precipitation events (Ontkean et
al., 2003). Semi-aquatic mammals and waterfowl that commonly occupy the wetland can also
contaminate wetland water and uplands with fecal matter (Hyer and Moyer, 2004; Kadlec et al.,
2007). Muskrats occupied the wetland as evidenced from newly constructed lodges some time
between June 25 and September 24, 2008.
6.3 Water quality characteristics of a newly constructed drainage
ditch
Construction of the drainage ditch at LR3 transported solutes previously stored in the
wetland downstream. Wetland water storage decreased exponentially when the drain was
completed. Within four hours ~80% of solutes and 30% of water exported had exited the
wetland. The newly constructed ditch acted primarily as a conduit, transporting solutes
downstream directly from the wetland. Although some solute concentrations were significantly
correlated with distance along the newly constructed ditch and had slopes that differed from the
Cl slope, these relationships were not consistent across time. For example, concentrations of
HCO3 were higher along the ditch length 4 hr and lower 6 hr since the start of the drainage
experiment and concentrations of orthoP were often constant along the length of the ditch with
the exception of the first point that differed. These data thus suggest that no consistent biotic or
abiotic processing occurred along the length of the new ditch. This result is in direct contrast
with previous studies that have shown that agricultural drainage ditches can act as solute sources
48
and/or sinks where changes in solute concentrations are attributed to sedimentation/resuspension,
adsorption/desorption, biotic uptake/release, and microbial mediated reactions such as
mineralization and nitrification (Skaggs et al., 1994; Sharpley et al., 2007; Strock et al., 2007).
However, the conditions reported in this study differed from existing studies by: i) the
experimental drainage ditch was shorter than ditches typically studied, such that the residence
time may be too short for processing to occur; ii) the ditch was new and lacked well established
aquatic vegetation and microbial communities; and iii) drainage occurred over a very short
period in late fall when water temperature was near freezing. Cold temperatures have been
shown by others to restrict nutrient uptake by vegetation, microbially mediated reactions, as well
as sorption and diffusion rates (Kadlec and Reddy, 2001). The drainage conditions studied here
are common in the prairies, and thus results given herein should be transferable to other
watersheds in the PPR, although rigorous quantitative testing is recommended.
Although there was no change in concentration along the length of the new ditch, there
was an increase in concentration of most solutes studied with time since the start of drainage.
This trend suggests that a vertical concentration gradient existed in the wetland such that the
water closest to the sediment and in the sediment porewater had the highest constituent
concentrations. Barica (1974), Fisher and Reddy (2001), and Barker et al. (2010) have all found
vertical concentration gradients and elevated concentrations in sediment pore water in similar
marshes and shallow lakes. Fecal coliforms also tend to concentrate in sediment where they
survive longer, potentially due to the greater organic matter present in the sediment than water
column (Karim et al., 2004). A vertical gradient in concentration may also explain why the
masses of TP, orthoP, NO3, NH4, E. coli and T. coli exported via the drainage ditch exceeded
estimates of those in the wetland. Wetland water samples were obtained from the center of the
wetland at half the water depth and thus the masses calculated from these samples would be
underestimated if a vertical concentration gradient existed.
Another explanation for the increase in concentration of water quality parameters is that
many wetlands in the prairies have salt rings around their perimeter. Salts are likely concentrated
by shallow groundwater fluxes at this transition zone between wetlands and their uplands.
Constructing a ditch that traverses this ring may have also contributed to the elevated
concentrations and mass exceedances observed at LR3. Further research is needed to determine
the cause of the mass exceedances because accurately estimating the amount of solutes leaving
the wetland is the most important factor for predicting downstream export and associated
ecological consequences, given that the ditch acted as a simple conduit.
Had it not been for the uncharacteristically high midsummer rainfall, constituent
concentrations leaving the wetland would probably have been higher and constituent mass
exported lower at the time of drainage. Proportions of nutrients transported by snowmelt runoff
typically exceed exports during rainfall events given that snowmelt runoff typically dominates
prairie hydrology (Timmons and Holt, 1977; Tiessen et al., 2010). Decreases in specific
conductivity and by association ion concentrations, have been attributed to rainfall and decreased
evaporation (Barica 1978; LaBaugh and Swanson, 2004). The increases in ion concentrations
between October 22 and the time of drainage were likely a result of their exclusion from the
overlying ice and their freezing-out into a reduced water volume (Barica, 1975; Schwartz et al.,
1978; Lilbaek and Pomeroy, 2008). In comparison, low coliform densities at the start of the
drainage experiment were likely the result of sedimentation on the wetland bed (Auer and
Niehaus, 1993; Wang and Doyle, 1998).
49
6.4 Comparing artificial ditches and natural spills
Similar to the results for the newly constructed drainage ditch, water quality was not
altered during transport along ditch or spill connections. Consequently, the significantly greater
concentrations of TDN, DOC, HCO3, K, and Ca observed in ditches than spills were most likely
due to differences in wetland water quality rather than differing physical attributes of drains and
spills. Ditching permanent and semipermanent wetlands effectively turns them into seasonal ones
because they are drained each year. Further, drained wetlands are predominantly located in
cropped areas, a situation typical across the PPR (Gunterspergen et al., 2002). In contrast, spills
flowed predominantly from permanent wetlands and were located only in grassland and wooded
areas. As outlined in section 4.1 of this report, seasonal wetlands are characterized by greater
concentrations of TDN and DOC, and K concentrations are greater in cropped wetlands than
grassland or wooded wetlands.
The ditches and spills also had significantly different physical characteristics. The ditches
were more channelized, longer, and had higher flow velocities. Although this had no apparent
direct influence on the quality of water moving along them, the differences are likely to influence
the impacts on downstream water bodies. Wetland drainage ditches are created to connect
wetlands to the watershed drainage network. In contrast, the short length of spills means they
often are transiently connecting wetlands to other wetlands. These different connection
characteristics suggest that wetland drainage has a higher likelihood of potentially enhancing
downstream nutrient, salt and bacteria loading than spills.
To date, the only study of prairie pothole drainage effects on downstream water quality
(nutrients only) was a modeling exercise that compared different scenarios of wetland restoration
at Broughton Creek watershed, Manitoba (Yang et al. 2008). Using the SWAT model, the
researchers ran a wetland restoration scenario in which the 2005 wetland area was increased to
match 1968 conditions. They predicted a 23% reduction in TN and TP loads to the stream using
an empirical nutrient export coefficient. In contrast, results presented herein suggest that a
nutrient export coefficient other than 1 is not warranted. This is likely the result of generally low
temperatures during the snowmelt period when ditches and spills flow and when new ditch
construction occurs. Fall ditch construction following harvest is more common than spring ditch
construction because wetter soils in spring impede access to wetlands by heavy machinery and
farmers are not preoccupied with seeding. Instead, the volume of water and mass of constituents
exported from wetland drainage will depend on: i) how effectively the ditch drains the wetland;
and ii) the water quality characteristics of the wetland that is influences by the combined
permanence and land use setting.
6.5 Stream water quality
There was a trend toward higher discharge for the subbasins with medium high and high
wetland drainage compared to the one with low wetland drainage. Higher discharge is
comparable to other studies that show wetland drainage increases effective contributing area and
stream discharge (Campbell and Johnson, 1975; Saskatchewan Watershed Authority, 2008; Yang
et al., 2008). Peak flows may have been delayed in the subbasin with low wetland drainage
because of increased water storage potential (Pomeroy et al., 2009; Fang et al., 2010).
The dominant anion switch from HCO3 to SO4 on the falling limb of the freshet
hydrograph coincident with an increase in SC to >520 μS/cm was apparent in all subbasins.
Therefore, it was not associated with wetland drainage. Instead, it likely occurred as proportion
50
of input from dilute snowmelt runoff decreased relative to that of saltier groundwater and more
permanent surface water bodies.
Subbasins characterized by moderately high and high wetland drainage had significantly
higher mean export coefficients of TP, orthoP, HCO3, SO4, Mg, and Ca than the low drainage
subbasin. These nutrients and salts have been shown in this report and by other researchers
(Waiser and Robarts, 2004) to be concentrated in isolated prairie potholes located in cropped
watersheds. The potholes most frequently drained in Smith Creek watershed are semi permanent
to permanent ones located in cropped areas.
The differences in water quality found among subbasins may have been partly attributed
to variations in the land use distribution. Effects of nonpoint source pollution on stream water
quality, particularly its eutrophic effect on downstream water bodies, have been widely
documented; see Carpenter et al. (1998) for a comprehensive review. However, it is more likely
that the differing mean export coefficients of TP, orthoP, HCO3, SO4, Mg, and Ca among
subbasins was related to wetland drainage degree rather than land cover variations. Guo et al. (in
press) report land covers in Smith Creek watershed as extracted from 2007 and 2008 classified
SPOT5 images. Their data show land covers were very similar among subbasins; albeit they
lumped subbasins SC4 and SC5 rather than reporting them separately (see Figure 4.3 for areas
covered by these basins).
Generally, no differences in the water quality parameters studied were found for
subbasins characterized by medium high and high wetland drainage. This is not surprising given
the relatively small difference (7%) in the proportion of wetlands lost for the two subbasins.
Unfortunately, site selection preceded wetland loss determination from remote sensing imagery.
It is recommended that future studies examine water quality in relation to stream flows across a
wider range of wetland drainage scenarios. Some subbasins of Smith Creek watershed have
much higher wetland loss than those studied here; for example, SC3 experienced ~81% wetland
loss between 1958 and 2007-2008. Total export of the various water quality constituents at Smith
Creek outlet, which is upstream of the town drainage (Langenburg Creek), was higher than the
sum of the three studied subbasins. This suggests the lower part of the watershed, where many
wetlands have been lost, is a significant source of nutrients, salts and bacteria to Smith Creek.
Future examination of water quality in the subbasins that make up the lower part of the Smith
Creek watershed is thus warranted.
It has been shown in other agricultural systems outside of the PPR that wetland restoration
reduces downstream nutrient loading (Woltemade, 2000). Water quality improvements are not
necessary linearly related to numbers or areas of wetlands gained as the cumulative effects of
wetland loss on downstream water quality are not usually additive and instead tend to be
nonlinear (Whigham et al., 1988). Stream flows are also important in determining how effective
wetlands are in maintaining good quality water downstream. For example, Johnston et al. (1990)
reported that wetlands removed more TP, ammonia and suspended solids at high flows but
removed more NO3 at lows. It would be very interesting to compare water quality in Smith
Creek under different wetland restoration scenarios.
6.6 Exceedance of Federal and Provincial water quality guidelines for
wetlands, ditches, spills and streams
TP concentrations in the newly constructed ditch, 6 of 7 ditches, and 3 of 5 spills
exceeded the Saskatchewan Watershed Authority (2007b) objective of 0.1 mg/L for the Lake
Stewardship Program water quality index. This objective was also exceeded in 40 of 67 wetlands
51
sampled. Exceedances occurred in 71%, 50%, and 59% of crop, grass, and wood wetlands, and
77%, 65%, and 40% of seasonal, semi-permanent, and permanent wetlands. The TP objective
was also exceeded for 45% of the LR3 wetland sampling instances. The TP objective was
exceeded for all samples from the streams draining the subbasins with medium high and high
wetland drainage and 63% of stream samples from the subbasin with low wetland drainage.
The majority of NO3 concentrations measured in the wetlands, newly constructed ditch,
ditches and spills, and streams did not exceed the Canadian Council of Ministers of the
Environment (CCME) Canadian Environmental Quality guideline for the protection of aquatic
life (2.9 mg/L; Canadian Council of Ministers of the Environment, 2003). However, this
guideline was exceeded twice in the LR3 wetland following snowmelt, in 2 of 7 ditches, and in 1
stream sample from the high wetland drainage subbasin. The CCME guideline for NH3 + NH4 is
greatly affected by temperature and pH; for typical conditions the guideline ranges from 18.5
mg/L (0oC, pH = 7.0) to 0.07 mg/L (15oC, pH = 9.0) (Canadian Council of Ministers of the
Environment, 2003). Concentrations of NH4 did not exceed these guidelines, with the exception
of the first two samples obtained following snowmelt in the LR3 wetland.
The use of water containing salt concentrations that exceed CCME guidelines for water
used for irrigation (100 – 700 mg Cl/L) and for livestock (1000 mg SO4/L and 1000 mg Ca/L) is
not recommended. Concentrations of Cl, SO4, and Ca measured in the 67 wetlands, ditches and
spills, and streams sampled in 2009 were below the CCME guidelines; thus drainage water
would not be anticipated to degrade water quality with regards to these parameters. The
guideline for SO4 was exceeded twice in the LR3 wetland when the wetland volume was lowest
and in the newly constructed drainage ditch 1 hour after the start of drainage.
The CCME indicator bacteria guideline for recreational water quality (200 E. coli per 100
ml) was only exceeded by the samples collected in the newly constructed ditch 23 hours after the
start of drainage. The CCME guideline for the protection of agricultural water for crop irrigation
(100 E. coli per 100 ml) was exceeded July 23, 2008 in the LR3 wetland. With the exception of
the November 20, 2008 sample in the LR3 wetland, all samples collected in the LR3 wetland and
the new ditch exceeded the animal-specific CCME guideline for livestock watering (2 E. coli per
100 ml of sample). E. coli densities exceeded the livestock watering guideline for 29%, 27% and
17% of the stream samples taken from subbasins with low, medium high, and high wetland
drainage, respectively
6.7 Stream ecosystem health
6.7.1 Effects of wetland drainage stress levels on receiving lotic
ecosystem health
Data from 2008 and 2009 show that wetland drainage stress does not have a significant
effect on the ecosystem health of the Smith Creek. However, the Langenburg Creek East of
Langenburg is significantly impaired relative to reference, as two of the three samples from this
site reveal significantly different assemblages (through CA Axis 1, Table 4.3) and one with
significantly impacted diversity. This impairment may be the result of stressors from the town of
Langenburg rather than wetland drainage as the samples are dominated by taxa such as
Nematoda, Ostracoda, Oligochaeta, and Chironomidae, which typically reflect high organic
pollution (Barbour et al., 1999). However, these results should merely be taken as an alarm that
52
warrants further study into the how drainage from the town of Langenburg may influence the
ecosystem health of Smith Creek.
The ditches included in this study did support aquatic benthic macroinvertebrate
assemblages, and in the instance of ditch DT 3 it even had healthy assemblage. The other three
ditches included in this study revealed stressed ecosystem health, particularly due to the richness
and diversity measures of their condition. The assemblages in these three stressed ditches were
dominated by the midges Chironomidae, the freshwater worms Oligochaeta and the biting midge
(or no-see-ums) Ceratopogonidae with few other species. These are all organisms tolerant to
environmental extremes and capable of multiple generations per year. Further, as the ditches dry
these organisms either die and leave resting eggs, burrow deep into wet areas, or emerge and
disperse as in the case of Chironomidae and Ceratopogonidae.
6.7.2 Findings on the influence of time on ecosystem health
Reassuringly, sampling efforts from 2008 whereby multiple samples from Smith Creek
North of Marchwell, Smith Creek East of Langenburg, and Langenburg Creek East of
Langenburg were taken all provided relatively consistent precision in conclusions about the
ecological condition at each site. As such, we are confident that single samples from each site in
this watershed are relatively accurate in estimating condition, and that am certain that
comparisons can be made through time both between years and between sampling periods within
a year. In particular, the site Smith Creek East of Langenburg from 2008 and SC 3 from 2009 are
overlapping sites and can be compared between the two years of study. The condition of this site
in the river did not change between years despite a few samples appearing “stressed” in 2008;
closer inspection shows that they are stressed because their diversity is higher than reference thus
can be interpreted as a positive measure of impact. Therefore, this mainstem Smith Creek site did
not vary in its ecosystem health between years.
The second temporal component of this study, comparing between collection periods in
the spring in 2009, produced results suggesting that ecosystem health may degrade as the stream
slows and dries. Specifically, our two most northerly sites, and thus lower order and ergo earlier
to cease flowing in the spring, dropped from a healthy assemblage of macroinvertebrates in the
early spring to impaired at SC 4 and stressed at SC5. Interestingly, SC 4 has medium high
drainage, while SC5 has low drainage. A potential explanation for the difference in impact
between these two sites is that although they are lower order, the medium high drainage of the
SC 4 site may have lead to a more flashy flow this spring while the low drainage of SC5 may
have provided for a more prolonged and sustained flow into the later spring. The two remaining
stream sites, TV1 and SC 3 are further downstream in the watershed and may have still had
adequate flows to support this healthy macroinvertebrate assemblage through into mid May.
Further investigation into the relationship between the flow in this watershed through time and
the timing of ecosystem health degradation is warranted based on these preliminary observations.
6.7.3 Unanswered aspects of drainage-biotic relationships and future
directions
The lack of baseline data makes it difficult to place sites such as this in perspective, and
such information would allow the monitoring of ecosystem health through time as drainage
practices in the watershed are either increased or decreased. Further, lacking is any information
on the effect drainage may have had on the wetland ecosystem health in the Smith Creek
watershed. Certainly completely draining a wetland and ploughing it into terrestrial agricultural
53
land will result in a complete loss of aquatic ecosystem health, but intermediate drainage to
control the expansion or merely to reduce the size of wetlands may have negative effects on the
benthic macroinvertebrates due to loss of habitat complexity, and loss of habitat quantity.
Developing a reference condition approach for wetlands would be extremely useful in setting
ecosystem health objectives for wetlands, and assessing the impacts of activities such as drainage
and land-use.
7 Conclusions and recommendations
7.1 Conclusions
Project results show that prairie wetlands act as traps for nutrients, ions and bacteria.
Temporally intensive measures of ions and DOC in one, permanent wetland suggest that
hydrology is a strongly regulator of their storage. In contrast, nutrient and bacteria storage in the
wetland appeared to be regulated by hydrological and biotic processes. Spatial variations in
wetland water quality can be attributed in part to different land cover and permanence classes.
Unexpectedly, there was no interactive effect of land cover and permanence classes on wetland
solute chemistry. Also demonstrated was that neither SC nor ion dominance can be used to
distinguish among wetland permanence classes at Smith Creek watershed. This lack of
association is in contrast with previous studies that have linked ion dominance patterns and SC
(as a proxy for net groundwater seepage rates) to wetland permanence. Overall, the results mean
knowledge of land cover and/or permanence class can be used to provide a reasonable estimate
of the water quality of a wetland.
Wetland water quality was found to be an important control of water quality in drainage
water. Thus, the occurrence of high nutrient concentrations (which were sometimes above
federal or provincial water quality guidelines) measured in the LR3 wetland at the onset of the
spring freshet has important implications because drained wetlands typically connect with
streams or other downstream water bodies at this time. Hence, wetland drainage may augment
downstream nutrient loads. Not assessed in the wetland drainage experiment was how changing
the permanent wetland to a temporary one will influence nutrient exports in future spring freshet
events. For instance, will nutrient concentrations remain high in the wetland from year to year,
and similar to those observed in seasonal wetlands at Smith Creek? Or, will nutrients become
progressively flushed from the stockpiles established in the drained wetland soils leading to
reduced nutrient exports over time?
Results also suggest that the efficiency with which a wetland is drained is an important
factor in quantifying downstream exports. The temporally intensive measures of drainage water
quality suggest that a vertical concentration gradient existed in the LR3 wetland such that the
water closest to the sediment and in the sediment porewater of the wetland had the highest
constituent concentrations. Thus wetland water samples obtained from the center of the wetland
at half the water depth are likely to provide an underestimate of the total solute mass stored in the
wetland if calculated from these samples.
Ditches had higher TDN, DOC, HCO3, K, and Ca than spills. Similar to the results for the
newly constructed drainage ditch, water quality was not altered during transport along ditch or
spill connections. Since water quality was not altered during transport along the length of ditch
or spill connections, differences in wetland water quality likely controlled ditch chemistry rather
than differing physical attributes of drains. Drained wetlands are predominantly located in
54
cropped areas, a situation typical across the PPR, whereas natural connections among wetlands
tended to be located in the grassland and wooded areas of Smith Creek watershed.
Nearly all water quality parameters studied were higher in streams draining subbasins
with greater wetland drainage. The provincial TP objective for protection of The low ratios of
DIN to DIP that indicate Smith Creek is usually N-limited at its outlet combined with the finding
that ditches export greater TDN than natural spills suggests that even small future increases in
wetland drainage in the watershed could lead to enhanced stream algal growth. As Smith Creek
is a sub-basin of the Assiniboine River, excess nutrient loadings from draining wetlands could
facilitate further eutrophication of Lake Winnipeg, into which the Assiniboine River drains. It is
important to remember that wetland drainage is not the only landscape stressor in watersheds
located in the Prairie Pothole Region that can impact downstream water quality.
This is the first to identify wetland drainage impacts on the instream ecosystem health.
Overall, these preliminary results indicate that wetland drainage does not have a dramatic impact
on the benthic macroinvertebrate assemblage. Due to the small number of sampling sites, results
of this work are best used to inform future hypotheses about the interaction between wetland
drainage and instream ecosystem health rather than being treated as a definitive conclusion with
which to base management of wetland drainage. Aspects of the hydrology such as reduced
permanence of wetlands produced by drainage could impact stream ecosystem health negatively
as flashier, less stable freshets may result in quicker drying and less diverse assemblages of
macroinvertebrates. Research is needed on this topic. Also lacking is any information about how
changes in the water level of the drained wetlands may have altered their benthic
macroinvertebrate assemblage or ecosystem health, and comparisons need to be made between
the positive gains in ecosystem health downstream receiving waters get due to drainage
compared to what may be lost from the wetlands being drained. It is hypothesized that the
increase in amount of water in a lotic site may be the cause of increased ecosystem health.
Conversely, decreasing water in a wetland via drainage could lead to declining wetland
ecosystem health if the benthic macroinvertebrates respond similarly (but inversely) in streams
and wetlands.
7.2 Recommendations
Recommendation 1:
Water quality along wetland drains was found to be similar to that in natural spills. Exceptions
are that TDN, DOC, HCO3, K, and Ca concentrations were found to be higher along wetland
drains. The important difference, however, is that spills typically connect wetlands to one
another whereas drains connect wetlands to streams. One option to consider that could reduce
downstream export of solutes and bacteria during wetland drainage is having wetland drains
emulate spills. Drains could be constructed in such a manner as to connect wetlands to other
wetlands so that solutes and bacteria remain stored in the watershed. It would be ideal to
consider locating these storage zones in areas of the watershed where land productivity is low.
Recommendation 2:
Little change in water quality along the length of wetland drains was found. This means that the
use of an empirical nutrient export coefficient is not required to be included in a model suited to
simulating spring freshet nutrient exports due to wetland drainage at Smith Creek. However,
knowing wetland water quality and how effectively the ditch drains the wetland are important to
55
predicting solute exports to streams. Future models should thus include representations of
wetland land use and permanence setting, and wetland drainage efficacy.
Recommendation 3:
Many wetland drains connect directly to roadside ditches that then empty into Smith Creek. An
evaluation of solutes losses or gains during transport along roadside ditches was not investigated
and future study is warranted.
Recommendation 4:
At Smith Creek, wetland drainage appears to strongly influence stream water quality, but not
stream biotic integrity. This conclusion has strong policy implications but has been determined
for only a small number of sites in one watershed in the prairies over a limited time period.
Basins in other parts of the prairies with differing wetland and drainage configurations, soils and
climate should be investigated to see if this conclusion holds elsewhere or is specific to Smith
Creek.
Recommendation 5:
The wetland drainage experiment performed showed that considerable solute and bacteria
exports occur during drainage of a permanent wetland. Needed is an examination of solute
release upon re-wetting of the wetland bed and subsequent export along drains for wetlands that
have experienced repeated draining.
Recommendation 6:
This study did not examine benthic macroinvertebrate assemblages in wetlands being drained.
As drainage reduces the permanence of a wetland, investigation into linkages between wetland
water availability and biotic health is needed.
8 Acknowledgements
We are grateful to the Saskatchewan Ministry of Agriculture for funding this project.
Additional graduate student support was received from an NSERC PGS-M. We thank Larisa
Barber, Deanne A. Schulz, Kevin Kirkham, Nicole Seitz, Logan Fang, Mike Solohub, and Adam
Minke for their help in the field. Two Smith Creek watershed residents, Barry Schultz and Don
Werle were instrumental in providing the logistical support needed to complete the field work.
Sorting assistance for the macroinvertebrate samples was provided by Janet Halpin. We thank
Adam Minke for producing several of the GIS maps and computing water storage in the LR3
wetland. Various archived GIS data were provided by Ducks Unlimited Canada, hydrometric
data for Smith Creek gauging station were provided by Water Survey of Canada, and most
taxonomic identifications were provided by AquaTax Consulting, Saskatoon. Discussions with
John Pomeroy and Kevin Shook were very important to the success of this project. This work
would not have been possible without the support and cooperation of the land owners and
producers of Smith Creek watershed, the Smith Creek Advisory Committee and the Langenburg
Regional Economic Development Authority.
56
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64
10 Other aspects
10.1 Communications
The research group utilizes several methods to communicate our project and its findings
with other researchers, the people living in the watersheds we study, and society. These include
conference talks, academic publications, public open houses and website development
(www.usask.ca/hydrology). Journal publications are forthcoming and we are currently preparing
an open house for the residents of Smith Creek watershed which is expected to be held in early
spring of 2011. Ms. Brunet will defend her thesis in March 2011. Below is a list of all scholarly
communications of this project thus far. CBC Saskatchewan Radio and Ducks Unlimited Canada
have asked for a copy of this report.
Brunet, N. and Westbrook, C.J. 2010. Prairie stream water quality in sub-basins characterized by
differing degrees of wetland drainage. American Geophysical Union 2010 Fall Meeting,
San Francisco, USA, December 13-17.
Brunet, N. and Westbrook, C.J. 2010. Prairie wetland drainage effects on water quality.
Canadian Soil Science Society Annual Meeting, Saskatoon, Canada, June 20-24.
Brunet, N. and Westbrook, C.J. 2010. Prairie wetland drainage effects on water quality. Joint
Assembly of the Canadian Geophysical Union and Canadian Meteorological and
Oceanographic Society, Ottawa, Canada, May 31 to June 4.
Brunet, N., and Westbrook C.J. 2010. Characterization of spatial variations in prairie wetland
water quality. CGU-HS Student Meeting, Edmonton, Canada, January 30.
Brunet, N. and Westbrook, C.J. 2009. Characterization of spatial variations in prairie wetland
water quality. Drought Research Initiative Conference, Saskatoon, Canada, November
18.
Westbrook, C.J. Pomeroy, J., Fang, X., Guo, X., Minke, A., Brunet, N., Shook, K., Brown, T.,
2009. Smith Creek and the importance of hydrometric data in modeling and water quality
research. Water Survey of Canada, Saskatchewan Division Staff Workshop, December,
Saskatoon, Saskatchewan.
10.2 Personnel involved
Cherie J. Westbrook – Principle Investigator, Assistant Professor, University of Saskatchewan
Iain Phillips – Aquatic Macroinvertebrate Ecologist, Saskatchewan Watershed Authority
John-Mark Davies – Water Quality Scientist, Saskatchewan Watershed Authority
Nathalie Brunet – MSc student, University of Saskatchewan
Erin Shaw – Lab technician, University of Saskatchewan
Summer Students - Larisa Barber, Deanne Schulz, Kevin Kirkham, Nicole Seitz
65
11 Appendices
Appendix A: Nutrient concentrations (May 19-21, 2009) for 67 wetlands in Smith Creek watershed, SK. Notes
indicate whether the upland area surrounding the wetland was tilled, used for grazing, and the type of crop harvested
in 2008, as reported by land owners.
ID
Land Cover
Notes
Permanence
Max Depth
orthoP
TP
TDN
NO3
NH4
DOC
(cm)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
DIN:DIP
-5.8
W3
Grass
Ungrazed
Seasonal
44
0.005
0.09
1.13
0.018
0.011
33.0
W4
Grass
Ungrazed
Semiperm
79
0.005
0.09
0.91
0.002
0.016
28.9
3.6
W6
Grass
Ungrazed
Semiperm
63
0.005
0.13
0.90
0.021
0.032
30.3
10.6
W10
Grass
Ungrazed
Seasonal
62
0.03
0.17
1.08
0.018
0.036
28.4
1.8
W11
Grass
Ungrazed
Permanent
102
0.01
0.12
1.39
0.013
0.038
36.5
5.1
W13
Grass
Ungrazed
Permanent
87
0.01
0.07
1.01
0.007
0.010
33.1
1.7
W17
Wood
Ungrazed
Permanent
54
0.14
0.24
1.17
0.045
0.038
33.1
0.6
W18
Wood
Ungrazed
Permanent
60
0.06
0.13
0.94
0.009
0.018
19.2
0.5
W21
Wood
Ungrazed
Seasonal
44
0.12
0.22
1.50
0.014
0.041
38.6
0.5
W22
Wood
Ungrazed
Permanent
58
0.18
0.29
1.35
0.006
0.041
32.6
0.3
W24
Wood
Ungrazed
Semiperm
56
0.35
0.43
1.09
0.013
0.014
33.9
0.1
W27
Wood
Ungrazed
Permanent
100
0.03
0.16
1.05
0.017
0.044
21.6
2.0
W28
Wood
Ungrazed
Permanent
96
0.01
0.11
1.12
0.008
0.014
27.8
2.2
W30
Wood
Ungrazed
Semiperm
59
0.47
0.53
1.30
0.007
0.017
35.0
0.1
W32
Wood
Ungrazed
Semiperm
52
0.02
0.22
0.99
0.002
0.018
24.2
1.0
W34
Crop
Canola/Tilled
Permanent
67
0.03
0.08
1.16
0.027
0.040
22.3
2.2
W35
Crop
Canola/Tilled
Permanent
76
0.1
0.18
1.29
0.035
0.025
26.2
0.6
W37
Crop
Canola/Tilled
Semiperm
78
0.03
0.17
1.31
0.025
0.014
30.9
1.3
W40
Crop
Canola/Tilled
Seasonal
34
0.005
0.19
1.49
0.002
0.014
37.1
3.2
4.0
W42
Crop
Canola/Tilled
Seasonal
48
0.005
0.55
0.88
0.006
0.014
26.8
W43
Crop
Canola/Tilled
Permanent
109
0.005
0.05
0.77
0.006
0.036
21.4
8.4
W45
Grass
Grazed
Semiperm
74
0.03
0.11
1.41
0.008
0.053
31.3
2.0
1.4
W46
Grass
Grazed
Permanent
98
0.03
0.08
1.10
0.008
0.035
26.3
W47
Grass
Grazed
Seasonal
36
0.04
0.18
1.13
0.023
0.012
30.9
0.9
W48
Wood
Grazed
Permanent
90
0.005
0.02
0.81
0.006
0.012
22.5
3.6
W50
Wood
Grazed
Seasonal
34
0.005
0.10
1.48
0.010
0.042
30.1
10.4
W54
Wood
Grazed
Semiperm
44
0.005
0.09
1.05
0.012
0.035
29.8
9.3
W55
Wood
Grazed
Permanent
70
0.005
0.02
0.88
0.007
0.013
24.0
3.9
W61
Wood
Grazed
Permanent
80
0.005
0.02
1.14
0.005
0.014
29.5
3.8
W67
Wood
Grazed
Semiperm
65
0.005
0.04
1.15
0.004
0.012
31.0
3.2
W68
Wood
Grazed
Seasonal
20
0.005
0.06
1.61
0.002
0.013
31.7
3.0
W69
Wood
Ungrazed
Seasonal
22
0.005
0.11
1.82
0.040
0.041
52.8
16.2
W71
Wood
Ungrazed
Semiperm
66
0.005
0.17
1.40
0.037
0.040
30.1
15.4
W72
Wood
Ungrazed
Seasonal
61
0.03
0.18
1.48
0.023
0.041
35.5
2.1
W73
Wood
Ungrazed
Semiperm
64
0.02
0.05
1.77
0.034
0.015
42.4
2.4
W75
Wood
Ungrazed
Seasonal
28
0.005
0.11
1.66
0.022
0.042
38.3
12.8
2.9
W85
Crop
Wheat/Tilled
Semiperm
62
0.01
0.19
1.23
0.005
0.024
30.2
W86
Crop
Wheat/Tilled
Permanent
102
0.03
0.09
1.86
0.015
0.014
51.5
1.0
W87
Crop
Wheat/Tilled
Seasonal
26
0.26
0.61
1.98
0.027
0.039
50.7
0.3
W88
Crop
Wheat/Tilled
Semiperm
70
0.3
0.93
0.95
0.222
0.029
24.2
0.8
W89
Crop
Wheat
Permanent
102
0.005
0.06
0.83
0.007
0.031
20.7
7.6
W90
Crop
Wheat
Semiperm
87
0.005
0.13
1.64
0.016
0.020
38.8
7.2
W91
Crop
Wheat
Seasonal
50
0.43
1.10
1.65
0.027
0.020
44.2
0.1
W93
Crop
Canola
Seasonal
74
0.02
0.11
1.34
0.036
0.002
33.0
1.9
W96
Grass
Ungrazed
Seasonal
44
0.005
0.10
1.74
0.022
0.013
44.9
7.0
W98
Grass
Ungrazed
Permanent
90
0.01
0.03
1.15
0.022
0.033
33.2
5.5
W100
Grass
Ungrazed
Seasonal
32
0.005
0.05
1.25
0.028
0.034
35.3
12.4
W101
Grass
Ungrazed
Permanent
66
0.005
0.03
0.87
0.004
0.034
27.5
7.6
W102
Grass
Ungrazed
Seasonal
22
0.03
0.87
2.49
0.007
0.025
55.3
1.1
66
Appendix A: continued
ID
W103
Land Cover
Grass
Notes
Grazed
Permanence
Max Depth
orthoP
TP
TDN
NO3
NH4
DOC
(cm)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
DIN:DIP
--
Permanent
97
0.12
0.29
1.00
0.002
0.045
29.7
0.4
W104
Grass
Grazed
Seasonal
26
0.03
0.22
1.41
0.005
0.009
33.6
0.5
W106
Grass
Grazed
Permanent
70
0.14
0.37
1.26
0.014
0.012
31.6
0.2
W107
Grass
Grazed
Permanent
98
0.005
0.07
0.79
0.019
0.023
24.8
8.4
W108
Grass
Grazed
Seasonal
20
0.01
0.42
2.77
0.026
0.048
49.6
7.4
W110
Grass
Grazed
Semiperm
46
0.005
0.12
0.85
0.226
0.015
24.2
48.1
W111
Crop
Canola
Semiperm
69
0.2
0.34
1.70
0.028
0.002
39.7
0.1
W112
Grass
Ungrazed
Semiperm
31
0.005
0.10
1.09
0.011
0.035
35.5
9.2
W113
Crop
Wheat
Permanent
91
0.005
0.05
1.06
0.045
0.035
27.5
16.0
W114
Crop
Wheat
Semiperm
84
0.38
2.80
1.60
0.008
0.039
37.3
0.1
W115
Crop
Wheat
Seasonal
34
0.13
1.30
1.83
0.022
0.016
42.0
0.3
W116
Crop
Canola
Permanent
80
0.005
0.12
1.34
0.016
0.084
37.5
20.0
W117
Crop
Wheat/Tilled
Seasonal
38
0.05
0.53
2.17
0.009
0.018
45.2
0.5
W118
Crop
Wheat/Tilled
Permanent
87
0.005
0.09
1.34
0.024
0.036
26.5
12.0
W119
Wood
Grazed
Permanent
70
0.01
0.08
1.31
0.226
0.015
34.7
24.1
W120
Grass
Grazed
Semiperm
74
0.005
0.07
0.90
0.014
0.015
28.1
5.8
W121
Grass
Grazed
Seasonal
55
0.02
0.57
0.90
0.025
0.036
19.6
3.0
W122
Grass
Grazed
Semiperm
60
0.005
0.08
0.92
0.026
0.033
24.2
11.8
67
Appendix B: Salt concentrations (May 19-21, 2009) for 67 wetlands in Smith Creek watershed.
ID
Land Cover
Permanence
pH
SC
Cl
HCO3
SO4
Na
Mg
Ca
K
(μS/cm)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
W3
Grass
Seasonal
6.75
163
2.1
89.8
0.8
1.3
9.7
8.6
16.0
W4
Grass
Semiperm
7.24
414
1.8
120.5
95.7
8.9
28.2
21.1
10.0
W6
Grass
Semiperm
7.31
656
1.7
124.7
210.0
19.3
68.5
39.2
11.9
W10
Grass
Seasonal
7.44
309
3.8
80.2
22.0
2.6
15.3
12.7
14.0
W11
Grass
Permanent
7.68
695
1.8
127.8
168.4
21.0
74.2
35.6
16.0
W13
Grass
Permanent
7.24
404
1.3
79.3
49.3
5.7
20.3
9.7
7.9
W17
Wood
Permanent
6.92
158
2.0
63.0
0.6
0.3
6.2
13.8
20.3
W18
Wood
Permanent
6.55
63
1.5
25.7
0.1
0.2
1.6
4.6
11.3
W21
Wood
Seasonal
6.78
149
2.3
54.4
6.3
0.3
4.8
11.3
23.4
W22
Wood
Permanent
6.86
131
1.9
62.8
0.5
0.2
4.8
10.0
18.9
W24
Wood
Semiperm
6.77
127
1.4
63.7
0.1
0.2
3.8
11.0
19.4
W27
Wood
Permanent
6.74
111
2.4
48.2
2.3
0.5
4.3
8.7
12.7
W28
Wood
Permanent
6.76
150
2.0
70.0
7.8
0.6
6.7
11.3
13.4
W30
Wood
Semiperm
6.83
127
1.7
64.7
0.1
0.2
3.8
9.8
21.5
W32
Wood
Semiperm
6.84
57
0.0
18.3
0.1
0.1
1.2
4.9
4.6
W34
Crop
Permanent
7.19
181
2.0
72.6
9.1
2.0
7.8
13.1
8.3
W35
Crop
Permanent
7.19
211
4.9
69.1
18.3
1.9
7.7
14.1
18.1
W37
Crop
Semiperm
7.26
296
3.3
95.3
40.0
3.8
10.0
16.3
9.6
W40
Crop
Seasonal
7.38
287
4.6
98.4
6.0
5.4
11.8
19.3
18.7
W42
Crop
Seasonal
6.80
147
9.5
56.2
1.0
2.3
4.8
11.5
13.4
W43
Crop
Permanent
7.59
230
2.9
66.9
2.7
1.7
6.6
10.4
11.2
W45
Grass
Semiperm
7.28
413
3.2
98.5
70.2
7.5
29.1
14.5
14.2
W46
Grass
Permanent
7.47
314
5.9
77.4
50.7
6.7
22.5
13.9
13.5
W47
Grass
Seasonal
7.10
414
12.4
103.4
52.1
6.9
25.4
22.3
12.8
W48
Wood
Permanent
7.75
306
2.0
113.2
39.7
4.5
17.5
18.3
12.9
W50
Wood
Seasonal
7.19
144
0.7
53.9
14.0
0.5
7.8
12.0
9.1
W54
Wood
Semiperm
8.62
451
1.9
62.7
114.6
9.9
41.2
30.7
17.3
W55
Wood
Permanent
8.42
374
0.7
66.7
44.3
2.4
14.3
19.6
7.3
W61
Wood
Permanent
7.74
422
3.9
94.6
81.7
6.8
24.5
18.2
18.2
W67
Wood
Semiperm
6.93
134
2.8
79.9
0.7
0.6
5.0
9.0
19.3
W68
Wood
Seasonal
6.88
112
1.2
55.5
1.3
0.4
4.7
10.8
12.6
W69
Wood
Seasonal
7.06
211
1.4
87.9
8.1
0.9
7.6
13.3
34.1
W71
Wood
Semiperm
7.09
182
2.0
80.4
3.9
0.8
5.9
11.3
16.7
W72
Wood
Seasonal
6.90
170
2.7
77.7
5.4
0.9
6.9
11.0
25.8
W73
Wood
Semiperm
6.94
196
1.7
110.9
9.2
1.5
10.3
12.7
22.5
W75
Wood
Seasonal
7.24
194
0.9
70.2
2.1
1.1
4.9
12.7
13.0
W85
Crop
Semiperm
7.07
504
3.5
125.5
111.0
4.2
27.7
32.6
18.9
W86
Crop
Permanent
7.91
577
6.1
93.3
140.8
10.6
38.4
23.3
23.9
W87
Crop
Seasonal
6.84
399
9.0
99.1
34.5
4.7
16.4
21.7
33.9
W88
Crop
Semiperm
7.19
677
7.0
135.4
167.4
11.0
57.8
40.3
42.6
W89
Crop
Permanent
8.16
1694
5.1
145.4
505.4
44.0
133.4
40.4
17.0
W90
Crop
Semiperm
7.51
506
3.4
68.7
92.6
6.1
20.1
27.0
12.6
W91
Crop
Seasonal
6.96
975
13.3
130.0
329.3
18.6
68.7
74.5
39.2
W93
Crop
Seasonal
7.09
391
6.8
115.1
31.9
3.2
16.1
24.0
33.2
W96
Grass
Seasonal
7.15
868
8.5
157.8
256.1
70.2
55.0
26.3
16.1
W98
Grass
Permanent
7.56
1241
8.6
161.6
417.9
120.1
107.8
53.0
20.9
W100
Grass
Seasonal
7.27
890
5.2
117.4
284.0
60.8
68.0
47.3
13.7
W101
Grass
Permanent
7.06
792
3.7
95.1
248.2
54.7
47.5
53.1
12.7
W102
Grass
Seasonal
7.80
865
5.8
212.1
177.6
24.8
54.5
34.3
19.6
68
Appendix B: continued
ID
Land Cover
Permanence
pH
SC
Cl
HCO3
SO4
Na
Mg
Ca
K
(μS/cm)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
W103
Grass
Permanent
7.98
584
7.6
114.1
108.7
12.7
43.5
27.1
18.9
W104
Grass
Seasonal
7.94
376
7.7
124.8
6.3
1.2
10.3
18.0
25.9
W106
Grass
Permanent
7.86
776
11.2
141.6
180.9
21.9
67.6
36.8
24.1
W107
Grass
Permanent
8.12
1303
13.6
160.1
464.5
80.7
128.9
31.1
23.5
W108
Grass
Seasonal
8.54
1557
46.3
116.8
742.6
100.9
166.1
135.9
32.2
W110
Grass
Semiperm
7.17
520
9.5
53.6
156.8
27.8
28.9
20.4
13.6
W111
Crop
Semiperm
7.37
795
15.1
139.5
191.4
12.1
47.0
49.8
55.8
W112
Grass
Semiperm
7.00
806
11.5
122.8
273.6
51.2
49.2
49.9
22.0
W113
Crop
Permanent
7.58
1396
10.1
163.1
538.6
61.7
137.1
60.3
24.3
W114
Crop
Semiperm
7.13
349
8.6
118.8
30.5
4.6
15.1
21.4
33.1
W115
Crop
Seasonal
7.08
314
11.6
131.6
8.8
1.5
8.3
19.6
39.1
W116
Crop
Permanent
8.05
1780
8.2
232.9
488.9
50.3
149.4
28.7
29.6
W117
Crop
Seasonal
7.23
771
4.2
162.0
187.6
6.6
52.9
71.4
22.2
W118
Crop
Permanent
7.43
823
1.7
103.6
329.0
20.0
89.9
76.5
16.2
W119
Wood
Permanent
8.24
307
1.5
165.5
7.7
2.9
19.5
15.2
16.5
W120
Grass
Semiperm
7.73
865
16.0
127.0
284.3
57.5
67.6
34.0
20.5
W121
Grass
Seasonal
6.59
98
2.0
33.7
2.9
0.6
3.7
8.3
9.0
W122
Grass
Semiperm
6.95
262
3.1
108.6
18.9
1.9
13.5
12.8
11.2
69
Appendix C: Taxa List Continued (Part 1 of 4)
Major Taxa To Family To Lowest Identification Amphipoda Gammaridae Gammarus lacustris Taltridae Hyalella azteca Hydrachnidia Platyhelminthes Rhynchobdellida Glossiphoniidae Helobdella elongata Anostraca Collembola Nematoda Oligochaeta Ostracoda Pelecypoda Sphaeriidae Pisidium Gastropoda Lymnaeidae Aplexa hypnorum Fossaria Fossaria/Stagnicola Pseudosuccinea Stagnicola Stagnicola elodes Physidae Aplexa hypnorum Physa Physa skinneri SWA 2008 16 92 6 52 334 185 2 2 15 106 7 30 1 1 SWA 2008 17 3 14 11 16 83 517 103 11 238 1 3 45 9 12 35 SWA 2008 18 5 5 2 8 1642 168 901 2 14 2 9 SWA 2008 19 7 182 1 SWA 2009 4 3 4 67 262 1 38 3 21 1 16 2 5 70
Smith Creek Watershed Site SWA SWA SWA SWA 2009 2009 2009 2009 5 6 7 8 402 4 4031 24 1 5368 2 2 17 5 59 730 20 64 4 5 45 4 9 17 10 14 4 2 1 8 2 6 5 SWA 2009 9 8 28 4 4 80 12 44 52 SWA 2009 10 4 146 6 102 1 SWA 2009 11 2 48 111 2 Total 3 414 4174 11 1 2 5516 30 1882 2402 1199 233 2 47 519 1 10 22 45 38 45 62 8 1 107 Appendix C: Taxa List Continued (Part 2 of 4)
Major Taxa To Family To Lowest Ientification Planorbidae Gyraulus Promenetus exacuous Coleoptera Dytiscidae Agabus Colymbetes Colymbetes exaratus Hydroporinae Hygrotus Hygrotus sellatus Ilybius Liodessus Neoporus Oreodytes Rhantus Rhantus sericans Haliplidae Haliplus Haliplus immaculicollis Hydraenidae Hydrophilidae Berosus Helophorus Paracymus Diptera Ceratopogonidae Atrichopogon Pericoma/Telmatoscopus Chaoboridae SWA 2008 16 4 6 1 9 3 4 150 SWA 2008 17 10 2 6 13 2 3 2 3 5 49 9 3 110 2 SWA 2008 18 13 13 2 21 5 105 SWA 2008 19 1 8 SWA 2009 4 27 4 11 1 6 2 2 1 2 2 1 26 10 24 71
Smith Creek Watershed Site SWA SWA SWA SWA 2009 2009 2009 2009 5 6 7 8 3 2 2 1 5 30 1 26 2 2 26 1 2 10 1 3 1 109 26 9 2 14 16 SWA 2009 9 24 4 76 104 SWA 2009 10 4 38 SWA 2009 11 1 39 Total 83 19 2 27 50 1 2 26 5 2 6 2 26 2 3 3 38 4 6 2 71 15 3 698 26 144 Appendix C: Taxa List Continued (Part 3 of 4)
Major Taxa To Family To Lowest Ientification Chaoborus Chironomidae Culicidae Aedes Dolichopodidae Empididae Muscidae Psychodidae Sciomyzidae Simuliidae Simulium Simulium venustum/verecundum Simulium vittatum Stratiomyidae Nemotelus Odontomyia Stratiomys Syrphidae Tabanidae Chrysops Tipulidae Tipula Ephemeroptera Baetidae Caenidae Caenis Hemiptera Corixidae Callicorixa audeni Cenocorixa dakotensis SWA 2008 16 497 114 9 1 10 10 3 58 10 1 2 2 41 1 SWA 2008 17 1415 26 6 7 2 16 135 279 120 17 56 1 1 6 203 1 17 2 SWA 2008 18 383 6 2 8 2 2 SWA 2008 19 115 3 40 55 1 SWA 2009 4 1 9 28 4 7 8 532 6 94 2 6 72
Smith Creek Watershed Site SWA SWA SWA SWA 2009 2009 2009 2009 5 6 7 8 26 2 1 1159 164 338 221 1 2 2 17 23 15 14 2 2 1 3 1 14 2 4 2 2 6 2 1 2 1 2136 23 17 17 2 1 SWA 2009 9 2 24 8 4 4 8 4 SWA 2009 10 56 2 SWA 2009 11 Total 30 24 4383 148 40 98 6 15 1 28 32 139 867 6 273 27 10 60 79 1 5 1 17 212 1 2182 79 5 1 Appendix C: Taxa List Continued (Part 4 of 4)
Major Taxa To Family To Lowest Ientification Hesperocorixa Hesperocorixa atopodonta Hesperocorixa laevigata Hesperocorixa michiganensis Hesperocorixa vulgaris Sigara bicoloripennis Sigara conocephela Sigara decoratella Sigara solensis Lepidoptera Odonata Anisoptera Libellulidae Zygoptera Coenagrionidae Enallagma/Coenagrion Lestidae Lestes Trichoptera Leptoceridae Oecetis Limnephilidae Anabolia bimaculata Asynarchus Limnephilus Philarctus/Limnephilus Grand Total SWA 2008 16 3 1 1 1 1 2 11 1 4 67 46 2 64 2 1993 SWA 2008 17 2 3 2 4 15 1 46 2 577 2 34 4 4349 SWA 2008 18 1 2 4 3327 SWA 2008 19 1 1 2 15 433 SWA 2009 4 1 1 2 6 2413 73
Smith Creek Watershed Site SWA SWA SWA SWA 2009 2009 2009 2009 5 6 7 8 2 2 2 5 2 2 4 26 6 27 4 2 27 1 2 1 4 14 3 100 1 12712 1263 428 70 SWA 2009 9 12 8 4 4 576 SWA 2009 10 2 1 421 SWA 2009 11 2 Total 5 3 1 3 2 9 2 6 13 17 38 20 43 30 119 66 3 4 760 5 34 4 15 27985 
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